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    INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY



    ENVIRONMENTAL HEALTH CRITERIA 88





    POLYCHLORINATED DIBENSO- PARA-DIOXINS AND DIBENZOFURANS













    This report contains the collective views of an international group of
    experts and does not necessarily represent the decisions or the stated
    policy of the United Nations Environment Programme, the International
    Labour Organisation, or the World Health Organization.

    Published under the joint sponsorship of
    the United Nations Environment Programme,
    the International Labour Organisation,
    and the World Health Organization

    World Health Orgnization
    Geneva


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    CONTENTS

    ENVIRONMENTAL HEALTH CRITERIA FOR POLYCHLORINATED
    DIBENZO-PARA-DIOXINS AND DIBENZOFURANS

    1. SUMMARY AND RECOMMENDATIONS

         1.1. Summary
               1.1.1. Sources
               1.1.2. Ambient levels and routes of exposure
               1.1.3. Toxicokinetics, biotransformation, and
                       biological monitoring
               1.1.4. Health effects
                       1.1.4.1   Animals
                       1.1.4.2   Humans
               1.1.5. Conclusion
         1.2. Recommendations

     2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES,
         ANALYTICAL METHODS
         2.1. Identity
         2.2. Physical and chemical properties
         2.3. Analytical methods
               2.3.1. General aspects
               2.3.2. Sampling strategy and sampling methods
               2.3.3. Extraction procedures
               2.3.4. Sample clean-up
               2.3.5. Isomer identification
               2.3.6. Quantification
               2.3.7. Confirmation
               2.3.8. Other analytical methods

    3. SOURCES OF ENVIRONMENTAL POLLUTION
         3.1. Production, synthesis, and use
         3.2. Industrial processes
         3.3. Contamination of commercial products
               3.3.1. Chlorophenoxyacetic acid herbicides
               3.3.2. Hexachlorophene
               3.3.3. Chlorophenols
               3.3.4. Polychlorinated biphenyls (PCBs)
               3.3.5. Chlorodiphenyl ether herbicides
               3.3.6. Hexachlorobenzene
               3.3.7. Rice oil
         3.4. Sources of heavy environmental pollution
               3.4.1. Industrial accidents
               3.4.2. Improper disposal of industrial waste
               3.4.3. Heavy use of chemicals
         3.5. Other sources of PCDDs and PCDFs in the
               environment
               3.5.1. Thermal degradation of technical
                       products
               3.5.2. Incineration of municipal waste

               3.5.3. Incineration of sewage sludge
               3.5.4. Incineration of hospital waste
               3.5.5. Incineration of hazardous waste
               3.5.6. Metal industry and metal treatment
                       industry
               3.5.7. Wire reclamation
               3.5.8. Traffic
               3.5.9. Fires and accidents in PCB-filled
                       electrical equipment
               3.5.10. Pulp and paper industry
               3.5.11. Incineration of coal, peat, and wood
               3.5.12. Inorganic chlorine precursors
               3.5.13. Photochemical processes
         3.6. Comparison of isomeric pattern and congener
               profiles from various sources

    4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND
         TRANSFORMATIONS

         4.1. Environmental transport
               4.1.1. Air
               4.1.2. Water
               4.1.3. Soil and sediments
         4.2. Environmental transformation
               4.2.1. Abiotic transformation
               4.2.2. Biotransformation and biodegradation
         4.3. Bioaccumulation
         4.4. Levels in biota
               4.4.1. Vegetation
               4.4.2. Aquatic organisms
               4.4.3. Terrestrial animals
               4.4.4. Human data
                       4.4.4.1   Adipose tissue
                       4.4.4.2   Blood plasma

    5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

         5.1. Air
         5.2. Water and leachate
         5.3. Soil and sediment
         5.4. Food
               5.4.1. Meat and bovine milk
               5.4.2. Human milk
               5.4.3. Rice
         5.5. Yusho and Yu-cheng episodes

    6. KINETICS AND METABOLISM OF 2,3,7,8-TETRACHLORODIBENZO-
         P-DIOXIN (TCDD) AND OTHER PCDDs

         6.1. Uptake, distribution, and excretion
               6.1.1. Studies on rats
               6.1.2. Studies on mice

               6.1.3. Studies on guinea-pigs
               6.1.4. Studies on hamsters
               6.1.5. Studies on monkeys
               6.1.6. Studies on dogs
               6.1.7. Studies on cows
               6.1.8. In vitro studies
         6.2. Metabolic transformation
               6.2.1. Studies on mammals
                       6.2.1.1   In vivo studies
                       6.2.1.2   In vitro studies
         6.3. Transfer via placenta and/or milk
         6.4. Matrix effects on the uptake
               ("bio-availability")

    7. EFFECTS OF TCDD AND OTHER PCDDs ON EXPERIMENTAL
         ANIMALS AND IN VITRO TEST SYSTEMS

         7.1. Acute toxicity
               7.1.1. In vivo studies on mammals
               7.1.2. In vitro studies on mammalian cells
               7.1.3. Studies on birds
               7.1.4. Toxicity of metabolites
               7.1.5. Modulation of the acute toxicity
         7.2. Short-term toxicity
               7.2.1. Studies on rats
               7.2.2. Studies on mice
               7.2.3. Studies on guinea-pigs
               7.2.4. Studies on hamsters
               7.2.5. Studies on monkeys
         7.3. Long-term toxicity
               7.3.1. Studies on rats
               7.3.2. Studies on mice
               7.3.3. Studies on monkeys
        7.4. Effects detected by special studies
               7.4.1. Wasting syndrome
               7.4.2. Hepatotoxicity
                       7.4.2.1   Morphological alterations
                       7.4.2.2   Hepatic plasma membrane
                                 function
                       7.4.2.3   Biliary excretion
               7.4.3. Porphyria
               7.4.4. Epidermal effects
                       7.4.4.1   In vivo studies
                       7.4.4.2   In vitro studies
               7.4.5. Effects on the immune system
                       7.4.5.1   Histopathology
                       7.4.5.2   Humoral-mediated immunity
                       7.4.5.3   Cell-mediated immunity
                       7.4.5.4   Macrophage function
               7.4.6. Myelotoxicity
               7.4.7. Effects on the intermediary
                       metabolism

               7.4.8. Enzyme induction
                       7.4.8.1   Studies on rats
                       7.4.8.2   Studies on mice
                       7.4.8.3   Studies on guinea-pigs
                       7.4.8.4   Studies on rabbits
                       7.4.8.5   Studies on hamsters
                       7.4.8.6   Studies on cows
                       7.4.8.7   Studies on chick embryos
                       7.4.8.8   Studies on cell cultures
               7.4.9. Endocrine effects
               7.4.10. Vitamin A storage
         7.5. Embryotoxicity and reproductive effects
               7.5.1. Studies on rats
               7.5.2. Studies on mice
               7.5.3. Studies on rabbits
               7.5.4. Studies on monkeys
               7.5.5. Studies on chickens
         7.6. Mutagenicity and related end-points
               7.6.1. Mutagenicity
                       7.6.1.1   Studies on bacteria
                       7.6.1.2   Studies on eukaryotic cells
                       7.6.1.3   In vivo studies
               7.6.2. Interaction with nucleic acids
               7.6.3. Cytogenetic effects
               7.6.4. Cell transformation
         7.7. Carcinogenicity
               7.7.1. Long-term animal studies on single
                       compounds
               7.7.2. Long-term animal studies with mixed
                       compounds
               7.7.3. Short-term and interaction studies
         7.8. Mechanisms of action
               7.8.1. Receptor-mediated effects
               7.8.2. Toxicokinetics
               7.8.3. Impairment of normal cellular regulatory
                       systems
                       7.8.3.1   Endocrine imbalance
                       7.8.3.2   Body weight regulation
                       7.8.3.3   Plasma membrane function
                       7.8.3.4   Impaired vitamin A storage
               7.8.4. Lipid peroxidation

    8. EFFECTS OF PCDDs ON HUMAN BEINGS - EPIDEMIOLOGICAL
         AND CASE STUDIES

         8.1. Occupational studies - historical perspective
         8.2. General population studies
               8.2.1. Missouri, USA
               8.2.2. Seveso, Italy
               8.2.3. Viet Nam
         8.3. Signs and symptoms in humans associated with
               TCDD exposure

               8.3.1. Skin manifestations
               8.3.2. Systemic effects
               8.3.3. Neurological effects
               8.3.4. Psychiatric effects
         8.4. Epidemiological studies
         8.5. Human experimental studies

     9. TOXICOKINETICS OF PCDFs

         9.1. Uptake, distribution, and excretion
               9.1.1. Studies with 2,3,7,8-tetrachlorodibenzo-
                       furan (2,3,7,8-TCDF)
               9.1.2. Studies with other PCDFs
         9.2. Metabolic transformation
         9.3. Transfer via placenta and/or milk

    10. EFFECTS OF PCDFs ON ANIMALS

         10.1. Acute toxicity
               10.1.1. Studies on rats
               10.1.2. Studies on mice
               10.1.3. Studies on guinea-pigs
               10.1.4. Studies on rabbits
               10.1.5. Studies on monkeys
         10.2. Short-term toxicity
               10.2.1. Studies on rats
               10.2.2. Studies on mice
               10.2.3. Studies on guinea-pigs
               10.2.4. Studies on rabbits
               10.2.5. Studies on hamsters
               10.2.6. Studies on monkeys
               10.2.7. Studies on chickens
         10.3. Chronic toxicity
               10.3.1. Studies on monkeys
         10.4. Effects detected by special studies
               10.4.1. Immunobiological effects
                       10.4.1.1   Histopathology
                       10.4.1.2   Humoral-mediated immunity
                       10.4.1.3   Cell-mediated immunity
               10.4.2. Enzyme induction
                       10.4.2.1  Studies on rats
                       10.4.2.2  Studies on mice
                       10.4.2.3  Studies on chickens
                       10.4.2.4  Studies on cell cultures
               10.4.3. Receptor binding
         10.5. Embryotoxicity and reproductive effects
         10.6. Mutagenicity
         10.7. Carcinogenicity

    11. EFFECTS OF PCDFs ON HUMAN BEINGS

         11.1. Yusho and Yu-cheng

    12. EVALUATION OF HEALTH RISKS FROM THE EXPOSURE TO
         CHLORINATED DIBENZO-P-DIOXINS (PCDDs) AND
         DIBENZOFURANS (PCDFs)

         12.1. Introduction
         12.2. Exposure assessment
               12.2.1. Sources of contamination
               12.2.2. Ambient levels
               12.2.3. Routes of exposure
               12.2.4. Bioavailability
         12.3. Animal data
               12.3.1. Toxicokinetics of 2,3,7,8-TCDD
               12.3.2. Toxicokinetics of PCDDs and PCDFs,
                       other than TCDD
               12.3.3. Toxic effects 2,3,7,8-TCDD
               12.3.4. Toxic effects of PCDDs and PCDFs,
                       other than TCDD
               12.3.5. Review of species differences
         12.4. Human health effects
               12.4.1. PCDDs
               12.4.2. PCDFs
               12.4.3. Human body burden and kinetics
         12.5. General conclusions

    13. RECOMMENDATIONS

    14. EVALUATIONS BY INTERNATIONAL BODIES AND THE CONCEPT
         OF TCDD EQUIVALENTS

         14.1. International evaluations
         14.2. Methodologies used in assessment of
               risk from PCDDs and PCDFs
               14.2.1. Individual congeners
               14.2.2. Mixtures of PCDD and PCDF congeners and
                       isomers - concept of TCDD toxic
                       equivalents

    REFERENCES

    FRENCH TRANSLATION OF SUMMARY, EVALUATION, AND
    RECOMMENDATIONS
    

    WHO TASK GROUP ON CHLORINATED DIBENZO-p-DIOXINS AND
    DIBENZOFURANS


    Members

    Dr U.G. Ahlborg, Unit of Toxicology, National Institute of
       Environmental Medicine, Stockholm, Sweden
    Dr J.S. Bellin, Office of Toxic Substances, US Environmental
       Protection Agency, Washington, DC, USA
    Dr B. Birmingham, Ministry of the Environment, Hazardous Contaminants
       Section, Toronto, Ontario, Canada
    Professor A.D. Dayan, Department of Health and Social Security,
       St Bartholomew's Hospital Medical College, London, United
       Kingdom (Chairman)
    Dr A. di Domenica, Instituto Superiore di Sanita, Rome, Italy
    Dr M. Greenberg, Department of Health and Social Security,
       Division of Toxicology and Environmental Protection, London,
       United Kingdom
    Dr R.D. Kimbrough, United States Department of Health and Human
       Services, Center for Disease Control, Atlanta, Georgia, USA
       (Now at the US Environmental Protection Agency Washington,
       DC, USA)
    Dr R. Koch, Department of Toxicology, Institute of Hygiene,
       Gera, DDR
    Professor C. Rappe, Department of Chemistry, University of
       Umea, Umea, Sweden
    Dr S. Safe, Texas A and M University, College Station, Texas,
       USA
    Dr H. Spielmann, Max von Pettenkofer Institute, Bundesgesundheitsamt,
       Berlin (West)
    Dr J. Vos, National Institute of Public Health and Environmental
       Hygiene, Bilthoven, Netherlands

    Representatives

    Dr A. Berlin, Health and Safety Directorate, Commission of the
       European Communities, Luxembourg
    Mrs E. Cox, Department of the Environment, London, United
       Kingdom
    Miss F.D. Pollitt, Department of the Environment, London,
       United Kingdom

    Secretariat

    Dr G.C. Becking, International Programme on Chemical Safety,
       World Health Organization, Research Triangle Park, North
       Carolina, USA (Secretary)

    Secretariat (contd)

    Dr H. Hakensson, Unit of Toxicology, National Institute of
       Environmental Medicine, Stockholm, Sweden (Temporary
       Adviser) (Rapporteur)
    Dr E. Johnson, International Agency for Research on Cancer,
       World Health Organization, Lyons, France
    Dr S. Tarkowski, Regional Office for Europe, World Health
       Organization, Copenhagen, Denmark

    NOTE TO READERS OF THE CRITERIA DOCUMENTS


         Every effort has been made to present information in the criteria
    documents as accurately as possible without unduly delaying their
    publication. In the interest of all users of the environmental health
    criteria documents, readers are kindly requested to communicate any
    errors that may have occurred to the Manager of the International
    Programme on Chemical Safety, World Health Organization, Geneva,
    Switzerland, in order that they may be included in corrigenda, which
    will appear in subsequent volumes.


                                 *    *    *


         A detailed data profile and a legal file can be obtained from the
    International Register of Potentially Toxic Chemicals, Palais des
    Nations, 1211 Geneva 10, Switzerland (Telephone No. 7988400 -
    7985850).

    ENVIRONMENTAL HEALTH CRITERIA FOR POLYCHLORINATED DIBENZO-PARA-
    DIOXINS AND DIBENZOFURANS

         A WHO Task Group on Environmental Health Criteria for
    Polychlorinated Dibenzo-para-dioxins and Dibenzofurans met at the
    Monitoring and Assessment Research Centre, London, United Kingdom,
    from 9 to 13 February, 1987. Dr M. Berlin opened the meeting and
    welcomed the members on behalf of the host Institute and on behalf of
    the United Kingdom Department of Health and Social Security, who
    sponsored the meeting. Dr G.C. Becking addressed the meeting on behalf
    of the three cooperating organizations of the IPCS (UNEP, ILO, and
    WHO). The Task Group reviewed and revised the draft criteria document
    and made an evaluation of the risks for human health and for the
    environment from exposure to polychlorinated dibenzo-p-dioxins and
    dibenzofurans.

         The drafts of this document were prepared by Dr U.G. Ahlborg, Dr
    H. Hakensson, and Dr B. Holmstedt, all of the National Institute of
    Environmental Medicine, Stockholm, Sweden, and by Professor C. Rappe
    of the University of Umea, Umea, Sweden.

         The efforts of all who helped in the preparation and finalization
    of the document are gratefully acknowledged.


                                 *    *    *


         Partial financial support for the publication of this criteria
    document was kindly provided by the United States Department of Health
    and Human Services, through a contract from the National Institute of
    Environmental Health Sciences, Research Triangle Park, North Carolina,
    USA - a WHO Collaborating Centre for Environmental Health Effects. The
    United Kingdom Department of Health and Social Security generously
    supported the cost of printing.

    ABBREVIATIONS


    AHH      aryl hydrocarbon hydroxylase
    ALA      aminolevulinic acid
    BGG      bovine gammaglobulin
    BHA      butylated hydroxyanisole
    BP       benzo(a)-pyrene
    CMI      cell-mediated immunity
    DEN      diethylnitrosamine
    diCDD    dichlorinated dibenzo-p-dioxin
    diCDF    dichlorinated dibenzofuran
    DMBA     dimethylbenzathraline
    ECOD     7-ethoxycoumarin-o-deethylase
    EGF      epidermal growth factor
    EH       epoxide hydratase
    EI       electron impact
    EROD     7-ethoxyresurofin-o-deethylase
    ETG      epidermal transglutaminase
    fg       femtogram (10-15g)
    GC       gas chromatography
    heptaCDD heptachlorinated dibenzo-p-dioxin
    heptaCDF heptachlorinated dibenzofuran
    hexaCDD  hexachlorinated dibenzo-p-dioxin
    hexaCDF  hexachlorinated dibenzofuran
    HMI      humoral-mediated immunity
    HPLC     high pressure liquid chromatography
    IARC     International Agency for Research on Cancer
    ip       intraperitoneal
    IR       infrared
    LOEL     lowest-observed-effect level
    MCPA     4-chloro-o-tolyloxyacetic acid
    MFO      mixed-function oxidase
    MS       mass spectrometry
    MSW      municipal solid waste
    ng       nanogram (10-9g)
    NMR      nuclear magnetic resonance
    NOEL     no-observed-effect level
    octaCDD  octachlorinated dibenzo-p-dioxin
    octaCDF  octachlorinated dibenzofuran
    PAH      polyaromatic hydrocarbons
    PCB      polychlorinated biphenyl
    PCDD     polychlorinated dibenzo-p-dioxin
    PCDF     polychlorinated dibenzofuran
    PCDPE    polychlorinated diphenylether
    PCPY     polychlorinated pyrene
    PCQ      polychlorinated quaterphenyl
    pentaCDD pentachlorinated dibenzo-p-dioxin
    pentaCDF pentachlorinated dibenzofuran
    pg       picogram (10-12g)
    SC       subcutaneous
    SCE      sister chromatid exchange

    SD       standard deviation
    SEM      standard error of the mean
    SIM      selected ion monitoring
    TCDD     2,3,7,8-tetrachlorinated dibenzo-p-dioxin
    TCDF     2,3,7,8-tetrachlorinated dibenzofuran
    TCP      trichlorophenol
    tetraCDD tetrachlorinated dibenzofuran
    tetraCDF tetrachlorinated dibenzofuran
    TPA      12-o-tetradecanoylphorbol-13-acetate
    triCDD   trichlorinated dibenzo-p-dioxin
    triCDF   trichlorinated dibenzofuran
    t3       triiodothyronine
    t4       thyroxine
    UDPGT    UDP-glucuronosyltransferase
    UV       ultraviolet
    2,4-D    2,4-dichlorophenoxyacetic acid
    2,4,5-T  2,4,5-trichlorophenoxyacetic acid
    3-MC     3-methylcholanthrene

    1.  SUMMARY AND RECOMMENDATIONS

    1.1  Summary

    1.1.1  Sources

         Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated
    dibenzofurans (PCDFs) are two series of tricyclic aromatic compounds
    with similar chemical and physical properties; they are ubiquitous in
    the environment. They do not occur naturally, nor are they
    intentionally produced. There are 75 positional isomers of PCDDs and
    135 isomers of PCDFs.

         The most important sources of contamination with PCDDs and PCDFs
    include:
         -    contaminated commercial chemical products, such as
              chlorinated phenols and their derivatives, and PCBs;
         -    incineration of municipal, hazardous, and hospital
              wastes, and of sewage sludges;
         -    automobile operation;
         -    fossil fuel combustion;
         -    overheating and emissions from fires involving PCBs;
         -    disposal of industrial wastes resulting from
              processes such as the production of chlorophenols and
              their derivatives, chlorophenol wood treatment, use
              of PCB fluids in electrical equipment, and wastes
              from pulp and paper processing.

    1.1.2  Ambient levels and routes of exposure

         The limited data available indicate that ambient levels of these
    compounds are very low in air, soil, and sediment, i.e. fg/m3 in
    air, ng/kg in soil and sediment. Levels of PCDDs and PCDFs up to 50
    ng/kg have been found in aquatic organisms in the general environment.
    Data on contamination of drinking water and commercial food are very
    limited.

         Exposure to these compounds in the general population probably
    occurs mainly through the food-chain.

         Some workers engaged in the production, use, and destruction of
    materials containing PCDDs and PCDFs and their precursors may receive
    high exposure. For these persons, inhalation and dermal contact are
    the primary exposure routes of concern.

    1.1.3  Toxicokinetics, biotransformation, and biological monitoring

         The bioavailability of PCDDs and PCDFs depends on the matrix they
    are in and the route of exposure. Data on bioavailability through
    inhalation are not available for any species.

         The quantity absorbed by humans after any route of exposure is
    not known.

         Studies on rodents given single or repeated oral doses of
    2,3,7,8-TCDD have shown that about half of the administered dose is
    absorbed from the gastrointestinal tract. The reported half-lives for
    elimination were between 12 and 94 days for rodents. The half-life of
    2,3,7,8-TCDD in adipose tissue of the rhesus monkey is about 1 year.

         Animal data on the toxicokinetics of PCDDs other than
    2,3,7,8-TCDD are limited. The half-life for 2,3,7,8-TCDD has been
    reported to be in the range of 2 and 8 days for rats, mice, and
    monkeys and more than 20 days for guinea-pigs. Studies on rats have
    shown that 2,3,4,7,8-pentaCDF is more highly retained than is
    2,3,7,8-TCDD.

         Data on the retention of PCDDs and PCDFs in tissues of various
    species, exposed to synthetic mixtures or to environmental samples
    containing PCDDs and PCDFs, show a high variability in retention time
    between congeners with or without chlorine substitution in the 2,3,7,
    and 8 positions.

         Limited human data indicate half-lives for some 2,3,7,8-
    substituted PCDDs and PCDFs in the range of 2-6 years.

         The PCDDs and PCDFs are predominately stored in fat, but they are
    also excreted in milk and pass through the placenta. They also appear
    in the blood and vital organs at lower concentrations.

         The tissue distribution in humans is not clear at present,
    although it has been suggested that the ratio between fatty tissue and
    liver is higher in humans than in rodents.

         In human fat, background levels of TCDD up to 20 ng/kg have been
    found in the general population, with no known specific exposure, but
    higher levels have been reported in some cases without evidence of
    disease. None of these populations were randomly sampled. The more
    highly chlorinated PCDDs and PCDFs, particularly octaCDD, are also
    present in these samples. Average tissue levels of TCCD tend to
    increase with age.

    1.1.4  Health effects

    1.1.4.1  Animals

         The toxic and biological effects resulting from exposure to
    2,3,7,8-TCDD are dependent on a number of factors, which include the
    species, strain, age, and sex of the animals used. The toxic responses
    observed in several animal species include body weight loss,
    hepatotoxicity, porphyria, dermal toxicity, gastric lesions, thymus

    atrophy and immunotoxicity, teratogenicity, reproductive effects, and
    carcinogenicity. TCDD induces a wide spectrum of biological effects
    including enzyme induction and vitamin A depletion. Not all of these
    effects are observed in any single animal species. The most
    characteristic toxic effects observed in all laboratory animals are
    body weight loss, thymus atrophy, and immunotoxicity. Chloracne and
    related dermal lesions are the most frequently noted signs of
    2,3,7,8-TCDD toxicosis in humans; dermal lesions are also observed in
    rhesus monkeys, hairless mice, and rabbits. In contrast, most rodents
    do not develop chloracne and related dermal toxic lesions after
    exposure to 2,3,7,8-TCDD. Many of the toxic lesions are noted
    primarily in epithelial tissues.

         Reproductive effects have been reported in rhesus monkeys and
    rats. The lowest-observed-effect levels have been reported to be
    approximately 1-2 ng/kg body weight per day. In two cancer studies in
    rats, hepatocellular carcinomas were produced at approximate dose
    levels of 0.1 g/kg body weight per day and 0.01 g/kg body weight per
    day. Doses of 0.001 g/kg body weight resulted in foci or areas of
    hepatocellular alteration. The incidence of certain hormone-dependent
    tumours was lower than in the control animals.

         TCDD does not appear to have mutagenic properties, and is
    therefore not likely to be genotoxic. Thus, it is assumed to be
    carcinogenic through an indirect mechanism.

         Several other PCDDs and PCDFs cause signs and symptoms similar to
    those of 2,3,7,8-TCDD, but there is a wide variation with regard to
    potency. There are 12 isomers that display higher toxicity, i.e., the
    tetra-, penta-, hexa-, and heptaCDDs and CDFs with four chlorine atoms
    in the symmetrical lateral positions 2,3,7, and 8. A mixture of two
    hexachlorodibenzo-p-dioxins (1,2,3,7,8,9- and 1,2,3,6,7,8-hexaCDD)
    has been demonstrated to possess carcinogenic properties in long-term
    animal studies, but at higher doses than those used in the study of
    TCDD. Dibenzo-p-dioxin and 2,7-diCDD failed to demonstrate
    carcinogenic properties. The relative toxic and biological potencies
    of PCDDs and PCDFs have been estimated using short-term studies in
    rats and mammalian cell cultures.

         There are marked species differences in the susceptibility of
    animals to the biological and toxic effects elicited by
    2,3,7,8-substituted PCDDs and PCDFs. For example, the oral LD50 values
    range from 0.6 g/kg body weight in guinea-pigs, to 5051 g/kg body
    weight in Golden Syrian hamsters for 2,3,7,8-TCDD. The tremendous
    variation in species and strain sensitivity to 2,3,7,8-TCDD and
    related compounds cannot be explained by the observed toxicokinetic
    differences. The toxicity and toxicokinetics of TCDD in monkeys most
    closely resemble the effects observed in humans. There is evidence in

    inbred mice that the cellular levels of the Ah receptor correlate, in
    part, with susceptibility to the biological and toxic effects of these
    compounds. The receptor has also been identified in other species
    including man. However, interspecies comparison of cellular Ah
    receptor levels do not explain fully the differences in sensitivity.

    1.1.4.2  Humans

         For occupational and accidental exposures to PCDDs and PCDFs, in
    spite of many clinical and follow-up studies, no clear-cut persistent
    systemic effects have been delineated except for chloracne. Other
    effects have been noted, but, apart from chloracne and perhaps minor
    functional disorders, none has been persistent.

         In some epidemiological studies of people exposed to a mixture of
    dioxins, furans, and other chemicals, an increased incidence of cancer
    at different sites has been claimed, but a number of factors limits
    confidence in the findings.

         In the Seveso accident, the only clear-cut adverse health effect
    recorded has been chloracne. Chloracne (193 cases) occurred in 1976
    and 1977, and 20 of those individuals still had active chloracne in
    1984. Many studies have been performed to find possible links between
    exposure to Agent Orange and health effects in civilians or military
    personnel in Viet Nam. However, the information available to date does
    not allow definite conclusions to be drawn with regard to effects on
    human reproduction or any other significant health effects.

         In the Missouri incident, children who showed acute illness when
    the contamination occurred in 1971 are now reportedly in good health.
    Furthermore, epidemiological studies in Missouri on populations
    exposed to lower concentrations of dioxins over longer periods of time
    have so far not revealed any significant health effects. Although no
    clinical symptoms were observed, there were indications of an effect
    on the cell-mediated immune system.

         The only documented intoxications with PCDFs in humans are the
    two instances of contamination of rice oil with PCDFs, PCBs, and PCQs,
    i.e., Yusho in Japan, 1968, and Yu-cheng in Taiwan, 1979. In total,
    several thousand people were acutely intoxicated. From the data it
    appears most likely that the causative agent was the PCDFs. The
    general symptomatology was similar to that seen in intoxications with
    TCDD, with the differences reflecting the intensity of exposure and
    the ages and sex of those exposed.

         The average daily intake of 2,3,7,8-substituted PCDFs by Yusho
    patients was estimated to be 0.1-0.2 g/kg body weight for a period of
    several months, while the lowest dose causing disease was estimated to
    be 0.05-0.1 g/kg body weight per day over a period of 30 days.

    1.1.5  Conclusion

         PCDDs and PCDFs occur throughout the environment and we all
    probably carry a body burden of them. They have sometimes produced
    complex toxic effects following occupational and accidental exposure.

         Based on the Yusho disease and experiments in sensitive species
    of monkeys, and making assumptions about the relative potencies of
    PCDDs and PCDFs, man and certain monkeys may have comparable
    sensitivity to these compounds. However, the uncertainties related to
    the real dose received by humans and the difficulties of assessing
    toxic effects other than chloracne in humans prevents a firm
    conclusion as to the relative resistance of humans to the toxic
    effects of these compounds. Exposure should be reduced to levels as
    low as reasonably practicable.

    1.2  Recommendations

    1.   Analytical interlaboratory validation and "round-robin" studies
    using standardized quality assurance and quality control procedures
    are needed to improve analytical methodology.

         Sampling strategy and analytical procedures and data
    interpretation should be optimized and standardized before undertaking
    surveys.

    2.   Further information is required about the origins and
    environmental distributon and fate of PCDDs and PCDFs.

         Further monitoring data, including time trends and determinations
    of isomer patterns, are required for environmental levels of PCDDs and
    PCDFs, especially for food, ambient air, and sediments.

    3.   Data should be obtained about the effects of PCDDs and PCDFs on
    environmental biota.

    4.   More information is required on the bioavailability of PCDDs and
    PCDFs from different matrices in the environment and from the diet.
    Exposure from these sources should be correlated with agricultural and
    industrial practices.

    5.   Simpler and less expensive chemical and biological methods
    suitable for screening for the presence of PCDDs and PCDFs should be
    developed and validated.

    6.   Studies to determine the mechanisms of toxicity of PCDDs and
    PCDFs are needed to support an evaluation of the differences in
    effects between species and to support an extrapolation to man.

    7.   Further investigation of immunotoxicity is important, including
    cytotoxic T-lymphocyte function. Studies of the effects of perinatal
    exposure and of the duration of actions on the immune system are
    important.

    8.   Long-term toxicity studies should be carried out, including
    multigeneration reproductive studies in different species with three
    of the most widespread PCDDs and PCDFs, namely 2,3,4,7,8-pentaCDF,
    1,2,3,7,8-pentaCDD, and octaCDD.

    9.   Because humans are exposed to complex mixtures of PCDDs and
    PCDFs, test systems, including human cell culture systems, should be
    developed further and validated for evaluating the toxic potency of
    these compounds and other mixtures. These systems can be used to study
    mechanisms of action, structure activity relationships, and
    interactive effects.

    10.  Investigations to examine the body burden and to correlate it
    with clinical effects and laboratory findings are indicated. Follow-up
    studies of previously exposed groups are important.

    2.  IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, ANALYTICAL METHODS

    2.1  Identity

         The polychlorinated dibenzo-para-dioxins (PCDDs) and
    polychlorinated dibenzofurans (PCDFs) are two series of almost planar
    tricyclic aromatic compounds with very similar chemical properties.
    The general formulae are given in Fig. 1.

     FIGURE 1

         The number of chlorine atoms can vary between 1 and 8. The term
    isomers refers to comparisons between compounds with the same
    empirical formulae. The term congeners refers to comparisons between
    compounds within the same series but with a different number of
    chlorine atoms. The number of positional isomers is quite large; in
    all there are 75 PCDDs and 135 PCDFs and the number of isomers for a
    certain number of chlorine atoms is given in Table 1.

         The nomenclature used in this document is based on the system
    used by Chemical Abstracts. The Chemical Abstracts System Registry
    Numbers (CAS RN) for a few PCDDs and PCDFs that have been cited in the
    literature are provided in Table 2.

    2.2  Physical and Chemical Properties

         A large number of the individual PCDDs have been synthesized by
    various methods and characterized, mainly by gas chromatography-mass
    spectrometry (GC/MS) (Buser & Rappe, 1980, 1984; Taylor et al., 1985;
    Rappe et al., 1985a) but also by using nuclear magnetic resonance
    (NMR) or ultraviolet (UV), infrared (IR), (Pohland & Yang, 1972; Kende
    et al., 1974), or X-ray analyses (Boer et al., 1973; Slonecker et al.,
    1983).

        Table 1.  Number of PCDD and PCDF isomers
                                                                         
              Number                  Number                    Number
        of chlorine atoms        of PCDD isomers          of PCDF isomers
                                                                         
               1                        2                         4
               2                       10                        16
               3                       14                        28
               4                       22                        38
               5                       14                        28
               6                       10                        16
               7                        2                         4
               8                        1                         1
                                       75                       135
                                                                         
    

        Table 2.  CAS RN for some PCDDs and PCDFs
                                                                               
         PCDD congener         CAS RN        PCDF congener          CAS RN
                                                                               
          2,3,7,8-TetraCDD    1746-01-6       2,3,7,8-TetraCDF     51207-31-9
        1,2,3,7,8-PentaCDD   40321-76-4     1,2,3,7,8-PentaCDF     57117-41-6
       1,2,3,6,7,8-HexaCDD   57653-85-7     2,3,4,7,8-PentaCDF     57117-31-4
       1,2,3,7,8,9-HexaCDD   19408-74-3    1,2,3,4,7,8-HexaCDF     70648-26-9
    1,2,3,4,6,7,8-HeptaCDD   35822-46-9    1,2,3,6,7,8-HexaCDF     57117-44-9
    1,2,3,4,7,8,9-HeptaCDD   58200-70-7    1,2,3,7,8,9-HexaCDF     72918-38-8
                   OctaCDD    3268-87-9    2,3,4,6,7,8-HexaCDF     60851-34-5
                                                                               
    
         Pyrolysis of chlorinated phenols yields small amounts of one or
    more PCDD isomers. Using this technique all the 22 tetraCDDs have been
    prepared (Nestrick et al., 1979; Buser & Rappe, 1980) as well as the
    14 pentaCDDs (Buser & Rappe, 1984) and 10 hexaCDDs (Lamparski &
    Nestrick, 1981; Buser & Rappe, 1984).

         Taylor et al. (1985) have synthesized, separated, and isolated
    all the 22 tetraCDD isomers. In Table 3 are listed some other isomers
    that have been synthesized and isolated.

         The most toxic and most extensively studied representative of the
    chlorinated dioxins (PCDDs) is
    2,3,7,8-tetrachlorodibenzo-para-dioxin (2,3,7,8-tetraCDD) (Fig. 2).
    It is commercially available, as are more than 10 other PCDD
    congeners.

         The empirical formulae, molecular weights, and some physical
    properties of a few PCDDs are given in Table 4.

        Table 3.  Synthetic method and melting point for some PCDDs
                                                                               

                 PCDD         Synthetic   Melting point        Reference
                Isomer         methoda       C
                                                                               
               1-Chloro-          1          80-90        Pohland & Yang, 1972
               2-Chloro-          1          88-89        Pohland & Yang, 1972
             1,3-Dichloro-        1       113.5-114.5     Kende et al., 1974  
             2,3-Dichloro-        1         163-164       Pohland & Yang, 1972
             2,7-Dichloro-        2         209-210       Pohland & Yang, 1972
             2,8-Dichloro-        3       150.5-151       Pohland & Yang, 1972
           1,2,4-Trichloro-       4         128-129       Pohland & Yang, 1972
           2,3,7-Trichloro-       1         157-158       Kende et al., 1974  
         2,3,7,8-Tetrachloro-     2         305-306       Pohland & Yang, 1972
         2,3,7,8-Tetrachloro-     5         305-307       Kende et al., 1974  
         1,2,3,4-Tetrachloro-     4         188-190       Pohland & Yang, 1972
         1,3,7,8-Tetrachloro-     1       193.5-195       Kende et al., 1974  
         1,3,6,8-Tetrachloro-     2         219-219.5     Pohland & Yang, 1972
       1,2,3,4,7-Pentachloro-     5         195-196       Kende et al., 1974  
     1,2,3,4,7,8-Hexachloro-      5         275           Pohland & Yang, 1972
     1,2,4,6,7,9-Hexachloro-      2         238-240       Pohland & Yang, 1972
                 Octachloro       2         330           Pohland & Yang, 1972
                                                                               


    a Synthetic methods as follows:

         1 = Catechol + chlorobenzene
         2 = Pyrolysis of chlorphenols
         3 = Cyclization of chlorophenoxyphenol
         4 = Catechol + chloronitrobenzene
         5 = Chlorination of chlorodibenzodioxin
    

    FIGURE 2

        Table 4.  Physical properties of some PCDDs
                                                                                  

                         Molecular    Molecular   Absorption
       Compound          formulae      weight      maximum     Reference
                                                 (chloroform)
                                                     (nm)
                                                                                  

       2,3,7,8-TCDD      C12H4Cl4O2     321.9        310       Pohland &
                                                               Yang (1972)

    1,2,3,7,8-PentaCDD   C12H3Cl5O2     356.5        308       Gray et al.
                                                               (1976)

    1,2,3,6,7,8-HexaCDD  C12H2Cl6O2     390.9        316       Gray et al.
                                                               (1975)

    1,2,3,7,8,9-HexaCDD  C12H2Cl6O2     390.9        317       Gray et al.
                                                               (1975)
                                                                               
    

         Although tetraCDD is lipophilic, it is only slightly soluble in
    most solvents and very slightly soluble in water (Table 5).


        Table 5.  Solubility of 2,3,7,8-tetraCDD in various solventsa
                                                                               
         Solvent                               Solubility at 25 C
                                          g/litre                g/kg          
                                                                               
     O-Dichlorobenzene                     1.8                   1.4
     Chlorobenzene                         0.8                   0.72
     Perchloroethylene                     0.68                  0.48
     Chloroform                            0.55                  0.37
     Benzene                               0.47                  0.57
     Acetone                               0.09                  0.11
     Dimethylsulfoxideb                  < 0.1                 < 0.1
     Methanol                              0.01                  0.01
     Water                               2 x 10-7             2 x 10-7
                                                                               


    a     From: Crummett & Stehl (1973).
    b     DMSO caused detector fouling and a better value could not be obtained.
    
        Table 6.  Water solubility of PCDDsa
                                                                               
            Compound                Water solubility (g/litre)
                                20.0 C                  40.0 C           
                                                                               
          1,3,6,8-TetraCDD      (3.20.2) x 10-7         (3.90.4) x 10-7
          1,2,3,7-TetraCDD      (4.30.1) x 10-7        (12.70.8) x 10-7
        1,2,3,4,7-PentaCDD      (1.20.1) x 10-7         (4.60.1) x 10-7
       1,2,3,4,7,8-HexaCDD      (4.40.1) x 10-9        (19.00.1) x 10-9
    1,2,3,4,6,7,8-HeptaCDD      (2.40.3) x 10-9         (6.30.2) x 10-9
                   OctaCDD      (0.40.1) x 10-9         (2.00.2) x 10-9
                                                                               

    a From: Friesen et al. (1985).
    
         Marple et al. (1986a) have reanalysed the water solubility of
    2,3,7,8-TCDD and found it to be considerably less (12.5-19.2
    ng/litre). The log water-octanol partition coefficient (Kow) has
    been determined as 6.64 by Marple et al. (1986b).

         Friesen et al. (1985) have determined the water solubility for
    some PCDDs other than the 2,3,7,8-TCDD compound and these are given in
    Table 6.

         Similarly Webster et al. (1985) have determined the log
    octanol-water partition coefficients for a number of PCDDs (Table 7).

         2,3,7,8-TetraCDD is considered to be a stable compound, but due
    to its extreme toxicity its chemistry has not been fully evaluated.
    However, it undergoes substitution reactions (Baughman, 1974) as well
    as photochemical dechlorination (Crosby et al., 1971; Crosby & Wong,
    1977; Gebefugi et al., 1977). Thermally it is very stable and rapid
    decomposition of 2,3,7,8-tetraCDD occurs only at temperatures above
    750 C (Stehl et al., 1973). The other PCDDs have been much less
    studied; however, octaCDD is completely destroyed by treatment with
    hot alkali (Albro & Corbett, 1977).

         The first synthesis of 2,3,7,8-tetraCDD was reported by
    Sandermann et al. (1957), who used catalytic chlorination of the
    unchlorinated dioxin. It has also been prepared in good yields by the
    dimerization of 2,4,5-trichlorophenol salts (Buu-Hoi et al., 1971b;
    Langer et al., 1973).

         In the PCDF series, Mazer et al. (1983) synthesized all the 38
    positional tetraCDF isomers. The products were mixtures of isomers,
    and each of these isomers could be identified. Later Bell & Gari,
    (1985) isolated and characterized all the 38 tetraCDFs, 28 pentaCDFs,
    and 16 hexaCDFs.

        Table 7.  Values for log Kow for some PCDDs from linear and quadratic plots

                                                                               

                                log Kow (linear)        log Kow (quadratic)
                                                                               

                             Waters       Waters      Waters        Waters
                            Bondapak     Bondapak    Bondapak      Bondapak
          Compound                    (Woodburn data)          (Woodburn data)
                                                                               

          Dibenzo-p-dioxin   4.26          4.01         4.34          4.17
                 1-MonoCDD   4.81          4.52         4.91          4.75
                 2-MonoCDD   5.33          5.00         5.45          5.29
                 2,7-DiCDD   6.27          5.86         6.39          6.17
              1,2,4-TriCDD   7.36          6.86         7.45          7.11
          1,2,3,7-TetraCDD   8.15          7.58         8.19          7.72
          1,2,3,4-TetraCDD   8.63          8.02         8.64          8.07
          1,3,6,8-TetraCDD   8.70          8.08         8.70          8.12
        1,2,3,4,7-PentaCDD   9.48          8.80         9.40          8.64
       1,2,3,4,7,8-HexaCDD  10.40          9.65        10.22          9.19
    1,2,3,4,6,7,8-HeptaCDD  11.38         10.55        11.05          9.69
                   OctaCDD  12.26         11.35        11.76         10.07
                                                                              

    a       From: Webster et al. (1985).
    

         Kuroki et al. (1984) have synthesized 51 congeners of PCDFs by a
    structure specific method from chlorophenols and chloronitrobenzenes
    or chlorophenols and chlorodiphenyls iodonium salts. The structures
    were confirmed by MS and NMR.

         Safe & Safe (1984) described the synthesis of 22 PCDF congeners
    resulting in quantities of 10-320 mg of purified product. They also
    reported NMR data on the compounds synthesized.

         Sarna et al. (1984) and Burkhard & Kuehl, (1986) have documented
    the octanol/water partition coefficients for some PCDFs (Table 8). The
    disagreement for OCDF arises because of uncertainties in the Kow
    values of reference compounds of high Kow. The partitioning of
    organic chemicals between lipid and water is an important determinant
    of the bioconcentration potential of a toxicant and has sometimes been
    effectively used as an indicator of the preferred degradative in in
    vivo pathways.

        Table 8.  The logarithm of the octanol/water partition coefficients (Kow) of some PCDFs using HPLC methods

                                                                               

           PCDF                 log Kow              Reference
                                                                               

    2,8-dichloro-                5.95             Sarna et al. 1984a
                                 5.30b           Burkhard & Kuehl, 1986c

    2,3,7,8-tetrachloro-         5.820.02        Burkhard & Kuehl, 1986c

    octachloro-                 13.37             Sarna et al. 1984a
                                 8.78             Burkhard & Kuehl, 1986c
                                                                               

    a     Quadratic equation treatment: Biorad Biosil (10 mm) data.
    b     Quadratic equation treatment: unspecified "microbore" HPLC column.
    c     Sarna et al. (1984) data recalculated from experimental data.
    

    2.3.  Analytical Methods

    2.3.1  General aspects

         The earliest reported method used to detect 2,3,7,8-tetraCDD was
    a rabbit skin test (Adams et al., 1941). Test samples were applied to
    the inner surface of the ear and to the shaven belly of albino
    rabbits, and inflammatory responses were observed. Subsequently, Jones
    & Krizek (1962) developed a test based on the recovery and weight of
    the keratin formed on the rabbit ear after application of a sample.
    These biological methods were non-specific as to isomers and not
    sufficiently sensitive to detect low levels of contamination.

         In the late 1960s and early 1970s, gas chromatographic methods
    were used for the quantification mainly of 2,3,7,8-tetraCDD in
    commercial 2,4,5-T formations. The detection level was normally in the
    range of g/g. These analyses were not isomer-specific and the results
    could not be confirmed. Ryhage (1964) solved the problem of combining
    a gas chromatograph with a mass spectrometer. During the 1970s and
    1980s, various types of mass spectrometer and gas chromatograph/mass
    spectrometer combinations were used in analytical work. Use of these
    more sophisticated instruments allowed for the development of
    isomer-specific and validated analyses for the tetraCDDs in the very
    late 1970s and for the other PCDDs and PCDFs in the early 1980s.

         A number of spectroscopic methods are available for the
    laboratory identification of 2,3,7,8-tetraCDD, but their use is highly
    restricted, with the exception of mass spectroscopy (MS). Data on
    X-ray, infra-red (IR), ultra-violet (UV), nuclear magnetic resonance
    (NMR), electron spin resonance (ESR), and mass spectra were obtained
    by Pohland & Yang (1972), Baughman (1974), and Slonecker et al.
    (1983).

         Because of the large number of isomers and congeners, and due to
    the extreme toxicity of some PCDD and PCDF isomers, highly sensitive
    and specific analytical techniques are required for the measurements.
    Detection limits for the analysis of environmental and human samples
    should be orders of magnitude lower than the usual detection levels
    required for pesticide analysis. A detection level of 1 pg or less
    might be required to measure 2,3,7,8-tetraCDD and the other toxic
    isomers in a 1-g environmental sample. Analyses at such low levels are
    complicated by the presence of a multitude of other interfering
    compounds and clean-up procedures are required.

         The mono-, di-, and trichloro congeners are not usually included
    in these analyses. Such compounds are considered to be much less toxic
    than the higher chlorinated congeners and are also much more volatile
    and losses may occur during clean-up.

         It should be mentioned that the level of sophistication needed in
    the analyses for PCDD and PCDFs will depend upon the objectives
    thereof. In cases where the objectives were primarily to screen
    samples to identify groups of PCDDs and/or PCDFs (in a qualitative or
    semiquantitative manner), routine assays and bioassays were adequate.
    In other instances, where the objective of the analysis was to
    quantify accurately specific PCDD and/or PCDF isomers in the samples,
    sophisticated analytical procedures were required. Clearly, both types
    of analyses can be useful, depending on the purpose for which the
    analytical results are to be used.

         Many analytical methods have been developed in recent years for
    the analysis of trace amounts of PCDDs and PCDFs in environmental
    samples, especially for 2,3,7,8-tetraCDD. The most specific of these
    methods are based on MS. There are many requirements to be met by such
    an analytical method, including representative sampling and
    appropriate storage, efficient extraction, high selectivity in the
    clean-up, high specificity in the gas chromatography, high sensitivity
    in the detection, safe and reliable quantification, good
    reproducibility, useful confirmatory information.

         Several review articles discussing methods of analyzing PCDDs and
    PCDFs have appeared (McKinney, 1978; Esposito et al., 1980; Rappe &
    Buser, 1980; Harless & Lewis, 1982; Karasek & Anuska, 1982; Tiernan,
    1983; Crummett et al., 1985). Most of the older methods have been
    critically reviewed by a panel of experts assembled by the National
    Research Council of Canada (1981).

    2.3.2  Sampling strategy and sampling methods

         The quality and utility of analytical data depend on the validity
    of the sample and the adequacy of the sampling program. The purpose of
    sampling is to obtain specimens that represent the situation being
    studied. Sampling plans may require that systematic samples be
    obtained at specified times and places, or simple random sampling may
    be necessary. Generally, the sample should be an unbiased
    representation of the environmental situation.

         All aspects of a sampling programme should be planned and
    documented in detail, and the expected relationship of the sampling
    protocol to the analytical result should be defined. A sampling
    programme should include reasons for choosing sampling sites, the
    number and type of samples, the timing of sample acquisition and the
    sampling equipment used. A detailed sampling procedure should include
    a description of the sampling situation, the sampling methodology,
    labelling of samples, field blank preparation, pretreatment
    procedures, and transportation and storage procedures.

         The quality assurance programme should include means to
    demonstrate that containers or storage procedures do not alter the
    qualitative or quantitative composition of the sample. Special
    transportation and storage procedures (refrigeration or exclusion of
    light) should be described, if they are required.

         Because environmental samples are typically heterogeneous, a
    sufficiently large number of samples (ten or more) must normally be
    analyzed to obtain meaningful data on chemical composition. The number
    of individual samples that should be analyzed will depend on the kind
    of information required by the investigation. If an average
    compositional value is required, a number of randomly selected
    individual samples may be obtained, combined, and blended to provide
    a homogeneous composite sample from which a sufficient number of
    subsamples could be analyzed. If composition profiles, time trends, or
    the variability of the sample population are of interest, many samples
    need to be collected and analyzed individually.

         If field blanks are not available, efforts should be made to
    obtain blank samples that best simulate a sample that does not contain
    the specific chemical. In addition, measurements should be made to
    ascertain whether, and to what extent, any reagent or solvent used may
    contribute to or interfere with the analytical results (laboratory and
    solvent blanks).

         The recovery tests are frequently used and necessary to evaluate
    the analytical methodology. Uncontaminated samples from control sites
    that have been spiked with the chemical of interest provide the best
    information because they simulate any matrix effect. When feasible,
    isotopically labelled (13C, 37Cl) chemicals spiked into the sample
    provide the greatest accuracy since they are subjected to the same
    matrix effects. The 13C- and 37Cl-labelled compounds can be used
    to validate:

    (a) sampling (sampling surrogate),
    (b) analytical pretreatment (clean-up surrogate),
    (c) quantification (internal standard).

         Very few laboratories in the world have access to and experience
    in working with these complicated analyses.

         In order to be able to compare data generated in different
    laboratories, the same quantitative standard compounds should be used.
    Interlaboratory calibrations or "round-robin" studies have been
    performed in very few cases.

    2.3.3  Extraction procedures

         In this step, the sample is homogenized or digested and extracted
    with a suitable solvent or solvent mixture to remove the bulk of the
    sample matrix and transfer the PCDD and PCDF residue into the solvent.
    Both the selection of the proper solvent and the method of extraction
    can be critical in obtaining a satisfactory recovery of PCDDs and
    PCDFs from the sample matrix.

         Many different procedures for the extraction of PCDDs/PCDFs from
    various samples are described. In some cases this involves digestion
    or destruction of the matrix. Some of these methods have been
    evaluated in the report from the National Research Council of Canada
    (1982), while other methods are discussed by Tiernan (1983).

         An interlaboratory "round-robin" study involving 13 laboratories
    was carried out to evaluate the reliability of data on
    2,3,7,8-tetraCDD in fish samples. No significant differences were
    found from methods differing in the digestion or extraction procedures
    (Ryan et al., 1983b).

         In a study described by Albro et al. (1985), eight different
    approaches were applied in eight laboratories to quantify four PCDDs
    (2,3,7,8-tetraCDD; 1,2,3,7,8-pentaCDD; 1,2,3,4,7,8-hexaCDD; and
    octaCDD) and three PCDFs (2,3,7,8-tetraCDF; 2,3,4,7,8-pentaCDF; and
    1,2,3,7,8,9-hexaCDF) in spiked samples of an extract from human
    adipose tissue. Levels of fortification, unknown to the participating
    laboratories, were in the 5-50 ng/kg range, except for octaCDD (up to

    500 ng/kg). The results indicated that most of the procedures tested
    gave a high degree of qualitative reliability. However, other methods
    were not so accurate, a large portion of the reported data consisting
    of false positives or false negatives.

         Lustenhouwer et al. (1980) studied the extraction of PCDDs and
    PCDFs from a fly ash sample. A dramatic difference was found between
    different solvents.

    2.3.4  Sample clean-up

         In the sample clean-up, the PCDDs and PCDFs present in the sample
    should be separated from a multitude of other co-extracted and
    possibly interfering compounds. The clean-up methods, normally three
    steps or more, can vary for different sample matrices. Two different
    procedural trends can be recognized:

         (a)  all PCDD and PCDF isomers can be analyzed in one
              single fraction by the containment enrichment
              procedure (Norstrom et al., 1982; Stalling et al.,
              1983; Tiernan, 1983; Rappe, 1984),

         (b)  specific isomers are analyzed in different fractions
              mainly after normal-phase and reverse-phase high
              pressure liquid chromatography (HPLC) separation
              (Lamparski et al., 1979; Niemann et al., 1983; Tosine
              et al., 1983).

         This latter method allows the identification of only a few PCDD
    isomers in each fraction, and is mainly used to monitor TCDD and a few
    other congeners. For a monitoring program of all PCDDs and PCDFs a
    more general method might be preferred.

         The method described by Stalling et al. (1983) was originally
    designed for the analyses of fish samples. In a "round-robin" study of
    fish samples it gave good results (Ryan et al., 1983b). This method
    has now been used for the clean-up of other biological samples like
    bird muscle, seal fat, turtle fat, and human adipose tissue - blood,
    liver, kidney, and milk (Rappe et al., 1983c; Nygren et al., 1986;
    Rappe et al., 1986b).

    2.3.5  Isomer identification

         The purified extracts are used directly for the final analyses
    with the aid of a gas-chromatograph/mass spectrometer (GC/MS) equipped
    with a glass capillary or a fused-silica column. The column leads
    directly into the ion source of the mass spectrometer, which operates
    either in the electron impact (EI) or the negative ion-chemical
    ionization (NCI) mode. In view of the large variation in toxicological

    and biological effects of the PCDD and PCDF isomers, it is imperative
    that the isomers, particularly those having high toxicity, be
    identified. For an unambiguous isomer identification it is necessary
    to have access to all analytical standards within a specific group of
    isomers, e.g. all the 22 tetraCDDs and all the 38 tetraCDFs. All the
    22 tetraCDDs have been prepared and, using a Silar 10c glass capillary
    column, the 2,3,7,8-tetraCDD can be separated from all the other 21
    tetra isomers (Buser & Rappe, 1980). Recently all the 14 pentaCDDs and
    the 10 hexaCDDs have been prepared. Using the Silar 10c column all the
    2,3,7,8- substituted isomers can be separated from all the other
    isomers (Buser & Rappe, 1984). The SP 2330 fused silica column can
    also be used for this separation (Rappe, 1984).

         In the PCDF series, Mazer et al. (1983) have synthesized all the
    38 positional tetraCDF isomers. The products were mixtures of isomers,
    and each of these isomers could be identified using both an SP 2330
    and an SE 54 capillary column. Later, Bell & Gara (1985) isolated and
    characterized all tetra-, penta- and hexaCDFs. The SP 2330 column can
    separate most of these isomers (Rappe, 1984). The 1,2,3,7,8-pentaCDF
    co-elutes with the 1,2,3,4,8-isomer and the 1,2,3,4,7,8- hexaCDF with
    the 1,2,3,4,7,9-isomer, but they can be separated on less polar
    columns like OV-17 and DB-5.

         A very limited number of investigations has been performed using
    these complete sets of synthetic standards.

    2.3.6  Quantification

         Mass selective detection (mass fragmentography) has been used to
    quantify trace amounts of PCDDs and PCDFs in the samples by
    selectively monitoring M, M + 2, and/or M + 4 ions (SIM). The
    quantification is based on peak area measurements and a comparison of
    these areas using either isotopically labelled internal standards
    (13C or 37Cl) or calibration curves of external standards. As a
    first approach, it has been generally assumed that with the MS
    quantification technique, all isomers of a particular congener of PCDD
    or PCDF (e.g. the tetrachloro-isomers) have the same response factors.
    However, an investigation of 13 well-defined tetraCDF isomers has
    shown a three-fold variation in response factors with the EI mode and
    up to a 20-fold variation with the negative ion-chemical ionization
    mode. For the higher chlorinated homologues (penta, hexa) the
    variation was found to be less (Rappe et al., 1983b).

         Fung et al. (1985) have studied the mass spectra of 26 PCDF
    congeners. They found that the EI spectra are not particularly isomer
    specific, while positive ion-chemical ionization spectra show a
    greater degree of isomer distinction.

    2.3.7  Confirmation

         Quality control and quality assurance programs help to assure
    that positive data reported actually refer to specific PCDDs and PCDFs
    (Kloepfer et al., 1983). To provide reliable data:

    (a)  isomer specificity must be demonstrated initially and verified
         daily,
    (b)  the retention time must equal (within 3 seconds) the retention
         time for the isotopically labelled congener,
    (c)  the signal to noise ratio must be 2.5:1 or higher,
    (d)  the chlorine cluster must be within  10% of the theoretical
         values, given in Table 9,
    (e)  correct fragments, e.g., M+-COCl ions, must be with correct
         chlorine clusters.

         For confirmation, mass spectroscopy is the best technique now
    available. The EI mass spectral properties of PCDFs and PCDD have been
    described (Buser, 1975). The molecular (M+) and fragment ions of
    PCDDs and PCDFs show the typical, expected clustering due to the
    chlorine isotopes (Table 9). The typical fragmentation is M-COCl+,
    which is a useful fragment to study.

         Buser & Rappe (1978) have shown that observation of low mass ions
    can be used for the identification of the substitution pattern of
    PCDDs, which can be defined as the number of chlorine atoms on each
    carbon ring of the dioxin molecular; the 2,3,7,8-isomer has a 2:2
    pattern while 1,2,3,4-tetraCDD has a 4:0 pattern. However, these low
    mass ions may not be observed in spectra from environmental or
    biological samples.

         In the negative ion-chemical ionization mode, the PCDFs have the
    base peak due to M-, and the fragmentation produces the unusual
    M--34 ions (uptake of H and loss of Cl). Fragmentation of PCDDs in
    this mode is more conventional via loss of Cl yielding M--35 ions
    (Buser et al., 1985).

         Using EI technique and a quadropole instrument, the detection
    limits are 1-10 pg for the tetrachloro compounds and up to 10-50 pg
    for the octachloro compounds using selected ion monitoring or multiple
    ion detection (SIM or MID). Full mass spectra require 0.1-1 ng of
    compound (Buser et al., 1985). High resolution instruments can improve
    the sensitivity by one order of magnitude.

         The negative ion-chemical ionization mode, using methane gas as
    reagent, gas provides extremely good sensitivity for all PCDFs (tetra-
    to octachloro- compounds) and for the higher chlorinated PCDDs (penta-
    to octaCDD). The detection limits are in the 10-100 fg (10-15g)
    range using SIR or MID, which is 1 to 2 orders of magnitude better
    than EI (Buser et al., 1985). However, the negative ion-chemical
    ionization mode has very poor sensitivity for 2,3,7,8- tetraCDD under
    these conditions.

         Using low resolution MS instruments, a series of interfering
    compounds has been identified (Table 10). Some of this interference
    can be eliminated using high resolution MS instruments operating at
    8000 - 10 000 daltons. However, compounds with the same empirical
    formulae cannot be separated by MS technique; they are normally
    eliminated during the clean-up or separated by the gas chromatography
    step.

    2.3.8  Other analytical methods

         Paasivirta et al. (1977) have shown that 2,3,7,8-tetraCDD can be
    detected down to the pg level using a glass capillary column and a
    63Ni electron-capture detector. Combined with efficient clean-up
    procedures, this method has shown to be useful down to a level of 9
    ppt (Niemann et al., 1983), although positive samples need
    confirmation by mass spectroscopy (MID, SIM).

         Other techniques, such as enzyme induction and radioimmunoassay
    have been described and discussed by Firestone (1978) and McKinney
    (1978). McKinney et al. (1982) have used the radioimmunoassay method
    for determining 2,3,7,8-tetraCDD in human fat, and found the reliable
    sensitivity at 95% confidence interval to be 100 pg per sample.

         An analytical method based on the keratonization response of
    epithelial cells in an in vitro system has been described by
    Gierthy & Crane (1985b). This method can be an assay for dioxin-like
    activity in environmental and biological samples. A positive response
    was found for 2,3,7,8-tetraCDD at a concentration of 10-11 mol/litre.



    
    Table 9.  Isotopic abundance ratio ("cluster") of polychlorinated dioxins and dibenzofurans

                                                                                                                             

    Number of
    chlorine             M          M + 2       M + 4       M + 6       M + 8       M + 10      M + 12      M + 14
    atoms
                                                                                                                             

       1               100.0        33.7
       2               100.0        66.1        11.3
       3               100.0        98.4        32.7        3.8
       4                76.4       100.0        49.4       11.0         1.0
       5                61.2       100.0        65.5       21.6         3.6        0.3
       6                51.1       100.0        81.7       35.8         8.9        1.2         0.1
       7                43.8       100.0        97.9       53.4        17.6        3.5         0.4
       8                33.7        87.6       100.0       65.3        26.8        7.0         1.2         0.1
                                                                                                                             

    Table 10.  List of molecular ions of polychlorinated compounds present in some human and environmental samples
    and possibly interfering in the mass spectral analysis of PCDFs and PCDDsa

                                                                                                                             

                                                      Molecular ions (m/z,m+,m-) (chlorination)
    Compounds       mono-     di-       tri-      tetra-    penta-    hexa-     hepta-   octa-     nona-     deca-


    PCDDs                                           320       354       388       422      456       -         -
    PCDFs                                           304       338       372       406      440       -         -

    PCBs                                            290       324       358       392      426       460       494
    PCNs                                            264       298       332       366      400       -         -
    PCTs                        298       332       366       400       434       468      502       536       570
    PCDPEsb                     238       272       306       340       374       408      442       476       510
    PCPYsc            36        270       304       338       372       406

                                                                                                                             


    a   From: Buser et al. (1985).
    b   PCDPEs: Polychlorinated diphenylethers.
    c   PCPYs: Polychlorinated pyrenes.
    


    3.  SOURCES OF ENVIRONMENTAL POLLUTION

    3.1  Production, Synthesis, and Use

         PCDDs and PCDFs are not produced commercially. These compounds
    are in fact formed as trace amounts of undesired impurities in the
    manufacture of other chemicals such as chlorinated phenols and their
    derivatives, chlorinated diphenyl ethers, and polychlorinated
    biphenyls (PCBs). There is no known technical use for the PCDDs and
    PCDFs.

         The amount of total PCDDs entering the Canadian environment/year
    has been estimated to be about 1500 kg, and 75% of this amount has
    been estimated to be due to octaCDD alone (National Research Council
    of Canada, 1981). There is no estimation of the amount of PCDFs
    entering the environment anywhere in the world.

         Although the polychlorinated dioxins and dibenzofurans are not
    commercially produced, most of these compounds have been synthesized
    for research purposes in small quantities according to the reactions
    discussed in section 2.

    3.2  Industrial Processes

         In addition to the synthetic methods mentioned in section 2,
    2,3,7,8-tetraCDD may be formed during the industrial preparation of
    2,4,5-trichlorophenol from 1,2,4,5-tetra-chlorobenzene. This
    substitution reaction takes place at about 180 C, and when the
    solvent is methanol, the pressure rises to about 7 KPa. The formation
    of TCDD is an unwanted side reaction which takes place when the
    reaction mixture is heated to 230-260 C (Milnes, 1971). This reaction
    is exothermic, so that even higher temperatures may be attained
    resulting in uncontrolled conditions.

         In some factories ethylene glycol is used as a solvent in order
    to avoid the high pressure. As already pointed out by Milnes (1971),
    however, use of this solvent requires special precautions because of
    the occurrence of a base-promoted polymerization of ethylene glycol
    and decomposition reactions that produce ethylene oxide. These
    reactions are also exothermic; they may start spontaneously at
    temperatures above 180 C and proceed rapidly and uncontrollably to
    result in the formation of relatively large amounts of TCDD.

         After most of the solvent has been removed, the reaction mixture
    is acidified; the 2,4,5-trichlorophenol can be separated from
    2,3,7,8-tetraCDD by one or two distillations, with the result that
    2,3,7,8-tetraCDD is concentrated in the still-bottom residues. Up to
    1 mg/g of 2,3,7,8-tetraCDD in such residues has been reported
    (Kimbrough et al., 1984). Improper disposal of such residues is
    discussed in sections 4.4.2 and 9.

         Most of the 2,4,5-trichlorophenol produced is used for the
    preparation of herbicides such as 2,4,5-T (including various esters
    and salts, and the bactericide hexachlorophene).

         PCDDs and PCDFs are both formed as by-products during the
    manufacture of chlorinated phenols (2,4-dichloro-, 2,4,6-trichloro-,
    2,3,4,6-tetrachloro- and pentachlorophenol). The commercial
    chlorophenols are produced by two processes, i.e., by chlorination of
    the phenol using various catalysts and by the alkaline hydrolysis of
    an appropriate chlorobenzene. Apparently both reactions can lead to
    the formation of PCDDs as well as PCDFs, and the level of
    contamination is normally much higher here than in the production of
    2,4,5-trichloro-phenol (see section 3.3).

         PCDDs and PCDFs are also formed during the preparation of
    chlorinated diphenyl ether herbicides (Yamagishi et al., 1981) and
    hexachlorobenzene (Villeneuve et al., 1974). A series of PCDFs are
    formed during the production of PCBs (see section 3.3).

         Production equipment is often used for the production of several
    different chemicals. In the manufacture of chemicals on such equipment
    previously contaminated by PCDDs and PCDFs, both the products and
    waste generated can be contaminated. Thus, manufactured
    2,4-dichlorophenoxyacetic esters (2,4-D), which otherwise should not
    be contaminated by 2,3,7,8-tetraCDD, did indeed contain this dioxin
    because the equipment used had been employed previously to produce
    2,4,5-T and had not been cleaned properly (Federal Register, 1980).

         It should be pointed out that the primary occurrence of TCDD in
    the environment is possibly related to the synthesis of
    2,4,5-trichlorophenol, the use of products prepared from this compound
    (Table 11), and to incinerations reactions. The occurrence of the
    other PCDDs and PCDFs is related to the synthesis and use of a variety
    of other products (Table 12), some of which are quite common.

         The other PCDDs and PCDFs are also formed in a variety of
    incineration reactions (see section 4.5).

    3.3  Contamination of Commercial Products

    3.3.1  Chlorophenoxyacetic acid herbicides

         Depending on the temperature control and purification efficiency,
    the levels of 2,3,7,8-tetraCDD in commercial products may vary
    greatly. For example, the levels of 2,3,7,8-tetraCDD in drums of the
    herbicide Agent Orange placed in storage in the USA and in the Pacific

    before 1970 were between 0.02 and 47 mg/g. More than 450 samples were
    analyzed in this study, and the mean value was 1.98 mg/g (Young et
    al., 1983). Since Agent Orange was formulated as a 1:1 mixture of the
    butyl esters of 2,4,5-T and 2,4-D, the levels of 2,3,7,8-tetraCDD in
    individual 2,4,5-T preparations manufactured and used in the 1960s
    could have been as high as 100 mg/g.

         In analyses using high-resolution GC-MS, Rappe et al. (1978a)
    have reported that in other samples of Agent Orange (as well as in
    European and the USA 2,4,5-T formulations from the 1950s and 1960s),
    2,3,7,8-tetraCDD was the dominating compound of this group of
    contaminants. Only minor amounts of other PCDDs and PCDFs could be
    found, primarily lower chlorinated PCDDs, in samples of Agent Orange.

         As a result of governmental regulations, efforts were made during
    the 1970s to minimize the formation of 2,3,7,8-tetraCDD during 2,4,5-T
    production, and now all producers claim that their products contain
    less than 0.1 g 2,3,7,8-tetraCDD/g of product (Rappe et al., 1978a).
    At present, the chloro-phenoxy herbicides are not the major source of
    PCDDs and PCDFs in the environment.

         Sixteen samples of 2,4-D esters and amine salts from Canada were
    analyzed for the presence of PCDDs. Eight out of nine esters and four
    out of seven amine salts were found to be contaminated, with the
    esters showing significantly higher levels (210-1752 ng/g) than the
    salts (20-278 ng/g). The tetraCDD observed was the 1,3,6,8-isomer, as
    verified by a synthetically prepared authentic standard (Cochrane et
    al., 1981). In other studies, it has been found that no tetraCDD other
    than the 1,3,6,8-isomer elutes in this window. Hagenmaier et al.
    (1986) has reported that, unexpectedly, a German 2,4-D formulation
    contained 6.8 ng of 2,3,7,8-tetraCDD/g.

    
    Table 11.  Some commercial products that may be contaminated with
    2,3,7,8-tetraCDD, depending on the method of preparation

                                                                               

    Common name                                Chemical name
                                                                               

    2,4,5-Ta                           2,4,5-Trichlorophenoxyacetic acid

    2,4,5-T estersa                    n-butyl-, butoxy ethyl-, and
                                       iso-octyl-esters of 2,4,5-
                                       trichlorophenoxyacetic acid

    2,4,5-T saltsa                     dimethylamine salts of 2,4,5-
                                       trichlorophenoxyacetic acid

    Fenoprop                           esters of 2-(2,4,5-trichlorophenoxy)-
                                       propanoic acid

    Erbon                              ethyl ester of 2-(2,4,5-trichloro-
                                       phenoxy)-2,2-dichloropropanoic acid

    2,4,5-Trichlorophenol              2,4,5-Trichlorophenol
    Fenochlorphos                      O,O-Dimethyl O-2,4,5-trichlorophenyl
                                       phosphonothioate

    Trichloronate                      O-Ethyl 0-2,4,5-trichlorophenyl
                                       ethylphosphonothioate

    Hexachlorophene/isobac 20          2,2'-Methylene-bis (3,4,6-trichloro-
                                       phenol)
                                                                               

    a  There are numerous trade names for this product.
    
        Table 12. Some commercial products which may be contaminated with PCDDs
    other than 2,3,7,8-tetraCDD, and with PCDFs, depending on the method of
    preparation
                                                                                     
    Common name                                Chemical name
                                                                                     
    Bifenox                            Methyl-5-2,4-dichlorophenoxy-2-nitrobenzoate

    Chloranil                          2,3,5,6-Tetrachloro-2,
                                       5-cyclo-hexadiene-1,4-dione.

    2,4-D (esters and salts)           2,4-Dichlorophenoxyacetic acid
                                       and esters and salts

    2,4-DB and salts                   2,4-Dichlorophenoxybutyric acid and
                                       salts

    Dicamba                            3,6-Dichloro-2-methoxybenzoic acid

    Dicamba, dimethylamine salt        3,6-Dichloro-2-methoxybenzoic acid,
                                       dimethylamine salt

    Dicapthon                          Phosphorothioic acid
                                       o-(2-chloro-4-nitrophenyl)
                                       o,o-dimethyl ester

    Dichlofenthion                     Phosphorothioic acid
                                       o-2,4-dichloro-phenyl
                                       o,o-dialkyl ester

    Disul sodium (sesone)              2,4-Dichlorophenoxyethyl sulfate,
                                       sodium salt

    2,4-DP                             2- 2,4-Dichlorophenoxy propionic acid

    HCB                                Hexachlorobenzene

    Nitrofen                           2,4-Dichlorophenyl-p-nitrophenyl
                                       ether

    PCP and salts                      Pentachlorophenol and salts

    PCB                                Polychlorinated biphenyls

    2,4,6-TCP                          2,4,6-Trichlorophenol and salts
                                       2,3,4,6-Tetrachlorophenol and salts

                                                                                     
    Common name                                Chemical name
                                                                                     

    CNP                                1,3,5-Trichloro-2-(4-nitrophenoxy)
                                       benzene

    NIP                                2,4-Dichloro-1-(4-nitrophenoxy)
                                       benzene

    X-52                               2,4-Dichloro-1-(3-methoxy-4-nitro-
                                       phenoxy) benzene
                                                                               
    
    3.3.2  Hexachlorophene

         The bactericide hexachlorophene is prepared from
    2,4,5-trichlorophenol, also the key intermediate in the production of
    2,4,5-T. Due to additional purification, the level of 2,3,7,8-tetraCDD
    in this product is usually < 0.03 mg/kg (Baughman, 1974). Ligon & May
    (1986) reported 0.0047 mg/kg of TCDD in one hexachlorophene sample.
    However, hexachlorophene also contains about 100 mg/kg of a
    hexachloroxanthene, the 1,2,4,6,8,9-substituted isomer (Gthe &
    Wachtmeister, 1972).

    3.3.3  Chlorophenols

         Chlorophenols have been used extensively since the 1950s as
    insecticides, fungicides, mold inhibitors, antiseptics, and
    disinfectants. In 1978 the annual world production was estimated to be
    approximately 200 000 tons. The most important use of 2,4,6-tri-,
    2,3,4,6-tetra-, and pentachlorophenol, and their salts, is for wood
    preservation. Pentachlorophenol is also used as a fungicide for slime
    control in the manufacture of paper pulp and for a variety of other
    purposes such as in cutting oils and fluids, for tanning leather, and
    in paint, glues, and outdoor textiles. 2,4-Di- and
    2,4,5-trichloro-phenol are used for the production of 2,4-D and
    2,4,5-T herbicides (phenoxy acids), and 2,4,5-trichlorophenol for the
    production of hexachlorophene.

         Chlorophenols are produced industrially either by direct
    chlorination of phenol or by hydrolysis of chlorobenzenes, the actual
    process used depending on the isomer desired. Chlorination of phenol
    yields 2,4-di-, 2,4,6-tri-, 2,3,4,6-tetra-, or pentachlorophenol,
    while hydrolysis of chlorobenzenes is mainly used for the production
    of 2,4,5-tri- and pentachlorophenol (Nilsson et al., 1978).
    Chlorophenols may contain a variety of by-products and contaminants,
    such as other chlorophenols, polychlorinated phenoxyphenols, and
    neutral compounds like polychlorinated benzene and diphenyl ethers

    (PCDPEs), PCDDs, and PCDFs. Some of these contaminants may also occur
    in chlorophenol derivatives like phenoxy acids, other pesticides, and
    hexachlorophene. The possible presence of PCDDs and PCDFs in
    commercial products is of special significance because of their
    extraordinary persistence and toxicological properties (see sections
    7-9). A scientific criteria document for chlorophenols and their
    impurities in the Canadian environment has been prepared by Jones
    (1981, 1984). Chlorophenols were estimated to be the major chemical
    sources of PCDDs and PCDFs in the Canadian environment (Sheffield,
    1985).

         Buser & Bosshardt (1976) reported on the results of a survey of
    the PCDD and PCDF contents of pentachlorophenol (PCP) and PCP-Na from
    commercial sources in Switzerland. From the results, a grouping of the
    samples into two series can be observed: a first series with generally
    low levels (hexaCDD <1 g/g) and a second series with much higher
    levels (hexaCDD >1 g/g) of PCDDs and PCDFs. Samples with high PCDD
    values had also high PCDF values. For most samples, the contents of
    the PCDF contaminants were in the order:

    tetra- = penta- < hexa- < hepta- < octaCDD/CDF.

    The ranges of the combined levels of PCDDs and PCDFs were 2-16 and
    1-26 g/g, respectively, for the first series of samples, and 120-500
    and 85-570 g/g, respectively, for the second series of samples. The
    levels of octaCDD and octaCDF were as high as 370 and 300 g/g,
    respectively.

         Some PCP-Na samples analyzed showed the unexpected presence of a
    tetraCDD (0.05-0.25 g/g), which was later identified by Buser & Rappe
    (1978) as the unusual 1,2,3,4-substituted isomer. Table 13 collects a
    number of relevant analyses of these chlorophenol formulations. The
    levels of PCDDs and PCDFs are higher than for the phenoxy-acetic acid
    herbicides.

         It has also been reported that several positional isomers of
    PCDDs and PCDFs are present in the chlorophenols. However,
    isomer-specific methods have not been used in most of these
    investigations, and more research is necessary to identify all the
    isomers present for a risk evaluation of these products.

         Miles et al. (1985) have analyzed PCP samples for hexaCDDs from
    five different manufacturers using an isomer-specific analytical
    method. The study included both free PCPs as well as the sodium salts.
    Total hexaCDDs in PCPs ranged from 0.66 to 38.5 mg/kg, while in the
    sodium salts levels of hexaCDDs between 1.55 and 16.3 mg/kg were
    found. The most abundant hexaCDD isomer found in the free PCPs was the
    1,2,3,6,7,8 isomer; however, in the sodium salts the 1,2,3,6,7,9- and
    1,2,3,6,8,9-hexaCDD pair was the most abundant.


    Table 13.  Levels of PCDDs and PCDFs in commercial chlorophenols (g/g)a

                                                                                
                  2,4,6-           2,3,4,6-            PCP            PCP
                  Trichlorophenol  Tetrachlorophenol   Sample A       Sample B
                                                                                
    TetraCDDs     < 0.1            < 0.1              < 0.1         < 0.1
    PentaCDDs     < 0.1            < 0.1              < 0.1         < 0.1
    HexaCDDs      < 1              < 1                < 1             2.5
    HeptaCDDs     < 1               10                  0.5          175
    OctaCDD       < 1                2                  4.3          500
    TetraCDFs       1.5              0.5               < 0.1          < 0.1
    PentaCDFs      17.5             10                 < 0.1          < 0.1
    HexaCDFs       36               70                   0.03         < 0.3
    HeptaCDFs       4.8             70                   0.5           19
    OctaCDF       < 1               10                  1.1           25
                                                                                

    a     From: Rappe et al. (1979).
    

         Hagenmaier & Brunner (1987) has reported that 2,3,7,8-tetraCDD
    can be found in commercial pentachlorphenol formulation at levels of
    0.21-0.56 ng/g, while Hagenmeyer & Brunner (1986) report that
    1,2,3,7,8-pentaCDD was found in pentachlorophenol and
    Na-pentachlorophenates in concentrations of 0.9-18 ng/g.

    3.3.4  Polychlorinated biphenyls (PCBs)

         Vos et al. (1970) were able to identify PCDFs (tetra- and
    pentaCDFs) in samples of European PCBs (Phenoclor DP-6 and Clophen A
    60) but not in a sample of Aroclor 1260. The toxic effects of these
    PCB products were found to parallel the levels of PCDFs present. Bowes
    et al. (1975) examined a series of Aroclors, as well as the samples of
    Aroclor 1260, Phenoclor DP-6, and Clophen A-60 previously analyzed by
    Vos et al. (1970). They used packed columns and very few standard
    compounds, and reported that the most abundant PCDFs had the same
    retention time as 2,3,7,8-tetraCDF and 2,3,4,7,8-pentaCDF. Using a
    complete set of PCDF standards and an isomer-specific analytical
    method, Rappe et al. (1985d) determined the levels of
    2,3,7,8-substituted PCDFs in commercial PCB products (see Table 14).

    3.3.5  Chlorodiphenyl ether herbicides

         In 1981, Yamagishi et al. reported on the occurrence of PCDDs and
    PCDFs in the commercial diphenyl ether herbicides
    1,3,5-trichloro-2-(4-nitrophenoxy) benzene (CNP),
    2,4-di-chloro-1-(4-nitrophenoxy)benzene (NIP), and
    2,4-dichloro-1-(3-methoxy-4-nitrophenoxy)benzene (X-52). The total
    tetraCDD found was 14.0 mg/kg in CNP, 0.38 mg/kg in NIP, and 0.03 in
    X-52. Very few synthetic standards were used, but the major tetraCDDs
    were identified as the 1,3,6,8- and 1,3,7,9-isomers, the expected
    impurities in the starting material 2,4,6-trichlorophenol. No
    2,3,7,8-tetraCDD could be detected in these samples. In all three
    herbicides, total tetraCDF was between 0.3 and 0.4 mg/kg.

    3.3.6  Hexachlorobenzene

         Hexachlorobenzene was used for the control of wheat bunt and
    fungi. Villeneuve et al. (1974), analyzing three commercial
    hexachlorobenzene preparations, identified octaCDD and hepta- and
    octaCDFs. The levels and identity of the heptaCDF isomers were not
    given. Great variation in levels of octaCDDs between the three samples
    (0.05-211.9 mg/kg) was noted, as well as in the level of octaCDF
    (0.35-58.3 mg/kg).

    3.3.7  Rice oil

         In 1968 more than 1500 people in southwest Japan were intoxicated
    by the consumption of a commercial rice oil accidentally contaminated
    by PCBs, PCDFs, and polychlorinated quarterphenyls (Masuda &
    Yoshimura, 1982; Masuda et al., 1985). In 1979 a similar episode
    occurred in central Taiwan, the number of people involved here
    approaching 2000 (Chen et al., 1980, 1981). Both these accidents have
    been referred to as Yusho episodes, but now the Taiwan episode has
    been renamed Yu-cheng (see section 5.4.4.4).

         The total level of PCDFs in the Japanese rice oil was reported to
    be 5 g/g (Nagayama et al., 1976) and 5.6 g/g (Buser et al., 1978d).
    For the rice oil from Taiwan, Chen et al. (1985) reported the PCDFs
    levels to be in the range 0.18-1.68 g/g.

         Buser et al. (1978) analyzed the Japanese rice oil using glass
    capillary columns. They found about 50-60 PCDF congeners and also
    reported that the 2,3,7,8-tetraCDF was the major isomer among the
    tetraCDFs. However, it was later shown that in this column system the
    2,3,4,8-tetraCDF co-elutes with the 2,3,7,8-isomer, and in fact the
    2,3,4,8-isomer was the main constituent in this peak (Chen & Hites,
    1983; Masuda et al., 1985). The 2,3,7,8-substituted congeners were
    estimated to account for 10-15% of the total amount of PCDFs (Buser et
    al., 1978).



    
    Table 14.  PCDFs in commercial PCBs (ng/g)a

                                                                                                                             
                        TRI-        TETRA-              PENTA-                          HEXA-                   HEPTA-
                                                                                                                      

                        Total    2378   Total   12348   23478   Total   123479  123678  123789  234678  Total   Total
         PCB-type                               12378                   123478


         Pyralene         700      53      630     10       T      35      ND      ND    ND      ND         ND     ND
         A1254             63      19     1400    690     490    4000    2500    2100   190     130     10 000    960
         A1260             10      13      110     48      56     260     500     120   190      27       1500   1300
         A30              500      35      573     14      28     160      50      59    ND      ND        220      T
         A40             1300     180     2600     96       8    1700      79      68    ND       T        310     ND
         A50             7400    3300   20 000    760    1100    8000     700     360    18      98       3100     75
         A60              770     840     6900   1100     990    8100    1600     330   170     330       6800   2000
         T64               47      23      360     97     122     840     520     390    58      41       2600    220
         Clophen C        710      54     1200     34      30     270      ND       T    ND      ND          T     ND


    a From: Rappe et al. (1985d).
    T  = traces.
    ND = not detected.
    


    3.4  Sources of Heavy Environmental Pollution

    3.4.1  Industrial accidents

         Several industrial accidents occurring during the production of
    2,4,5-trichlorophenol have been described in the literature. In most
    of these accidents the pollution of 2,3,7,8-tetraCDD has been to
    factories with circumscribed occupational exposure (section 9).
    However, on 10 July, 1976, a runaway reaction in a factory at Meda
    near Seveso in Northern Italy resulted in the escape of a chemical
    cloud of trichlorophenol/phenate containing 2,3,7,8-tetraCDD.

         The cloud initially covered an area outside the factory 5 km long
    and 700 m wide. On the basis of the TCDD levels found in the
    contaminated soil samples, it has been estimated that 2-3 kg of TCDD
    was released in this accident. About 80% of this amount was deposited
    in an area of 15 ha, within a distance of about 500 m from the plant.
    The levels of soil contamination in three zones are given in Table 15
    (Pocchiari, 1978).

    3.4.2  Improper disposal of industrial waste

         In 1973, three horse arenas in Missouri, USA, were found to be
    contaminated by high levels of 2,3,7,8-tetraCDD; the highest value
    reported was about 30 g/g of soil (Kimbrough et al., 1977). This
    contamination resulted from the application, in 1971, of contaminated
    waste oil to control dust at these locations. The TCDD had originated
    at a hexachlorophene-producing factory in Verona, Missouri. Additional
    tri- and tetraCDDs were also found, but the major component was
    1,2,4,6,8,9-hexachloroxanthene, a compound which apparently can serve
    as a marker for this type of contamination. The xanthene is a normal
    by-product of hexachlorophene production and has never been associated
    with the production of 2,4,5-tri-chlorophenol or 2,4,5-T derivatives
    (Viswanathan & Kloepfer, 1986).

         In 1982, numerous sites of potential 2,3,7,8-tetraCDD
    contamination were discovered in eastern Missouri. The contamination
    originated from the same waste oil from the factory in Verona. The
    streets of the entire town of Times Beach, Missouri, had been sprayed.
    More than 10 000 soil samples from Missouri were analyzed. In this
    state more than 40 hazardous waste sites containing 2,3,7,8-tetraCDD
    were identified. Most of these contaminated sites resulted from the
    disposal of waste from the same factory in Verona. The highest level
    reported in these soil samples was 9648 mg TCDD/g (Viswanathan &
    Kloepfer, 1986).

         Another location of great concern is Love Canal, Niagara Falls,
    USA. Here, Smith et al. (1983) found high levels of 2,3,7,8-tetraCDD
    in storm sewer sediments taken from around the Love Canal waste
    disposal site. The highest value was 312 ng/g sediment.


 

    Table 15. Distribution of TCDD contamination in the Seveso area on the basis of soil sample analysesa

                                                                               
         Range                         Number of soil samples
        (g/m2)                                                                

                             Zone A      Zone B    Surrounding monitored area
                                                                               

        < 0.750                32          25                 249
          0.750  - 4.99        32          53                 128
          5.0  -  14.99         6          19                   2
         15.0  -  49.99        18           6                   0
         50.0  - 499.99        31           0                   0
        500.0 - 4999.99        18           0                   0
        > 5000                  3           0                   0

                                                                               

    a From: Pocchiari (1978).
         Zone A: high-level contamination, about 115 ha.
         Zone B: low level contamination, about 255 ha.
         Surrounding area: about 1400 ha.
    
    3.4.3  Heavy use of chemicals

         The Eglin Air Force Base in Northwest Florida, USA, has been used
    for the development and testing of aerial spraying equipment for
    military defoliation operations. During the period 1962-1970, a
    3-km2 test area was sprayed with 73 tons of 2,4,5-T. Analyses of
    archived samples of the formulations indicated that approximately 2.8
    kg of 2,3,7,8-tetraCDD had been applied as a contaminant of the
    herbicide. However, one 37-ha test grid received 2.6 kg of this TCDD
    from 1962 to 1964. Levels of 10-1500 ng/kg were found in 22 soil
    samples (the top 15 cm) collected and analyzed 14 years after the last
    application of herbicide to this site (Young, 1983).

    3.5  Other Sources of PCDDs and PCDFs in the Environment

    3.5.1  Thermal degradation of technical products

         The formation of 2,3,7,8-tetraCDD as a result of thermal
    reactions of 2,4,5-T and 2,4,5-T derivatives has been the subject of
    controversy. Heating 2,4,5-T salts at 400-450 C for 30 minutes or
    longer yielded approximately 1 g of 2,3,7,8-tetraCDD per kg of 2,4,5-T
    salt, while no TCDD was identified from the same treatment of 2,4,5-T
    acid or esters (Langer et al., 1973; Baughman, 1974). Using a more
    sensitive analytical method, Ahling et al. (1977) reported that 0.2-3
    mg of 2,3,7,8-tetraCDD was formed per kg of 2,4,5-T esters during

    combustion at 500-850 C. Two reports (Stehl & Lamparski, 1977;
    Andersson et al., 1978) have shown that 2,3,7,8-tetraCDD could not be
    found after burning samples of spiked or sprayed vegetation at 600 C.
    The combustion gases, soot, particles, and ashes were analyzed and the
    detection limit was 4 mg of TCDD/kg 2,4,5-T burned.

         Rappe (1978b) have studied the burning of material impregnated
    with various salts of chlorophenols. Very carefully purified
    2,4,6-tri- and pentachlorophenate were studied, in addition to a
    commercial formulation of 2,3,4,6-tetra-chlorophenate. The analytical
    method used in this study was not isomer specific, but the following
    conclusions can be drawn concerning the formation of PCDDs by thermal
    reactions:

         (a)  the expected dimerization products and the products formed
              in the "Smiles rearrangement" are the major PCDDs;
         (b)  no other thermal isomerization of the PCDDs formed can be
              observed;
         (c)  no formation of higher chlorinated PCDDs can be observed;
         (d)  octaCDD and other higher chlorinated PCDDs yield lower
              chlorinated dioxins in a nonspecific dechlorination
              reaction;
         (e)  a series of PCDFs was also observed.

         It has been found that PCBs can be converted to PCDFs under
    pyrolytic conditions. The pyrolysis of commercial PCBs in sealed
    quartz ampoules in the presence of air yielded about 30 major, and 
    more than 30 minor, PCDFs. The optimal yield of PCDFs was about 10%,
    calculated on the amount of PCB decomposed. Thus, uncontrolled burning
    of PCBs can be an important environmental source of hazardous PCDFs.
    Therefore, it was recommended (Buser et al., 1978a, 1978d) that all
    destruction of PCB-contaminated waste using incinerators must be
    carefully controlled. In the temperature range 300-400 C, the
    conversion yield seems to be in the part-per-million range (Morita et
    al., 1978).

         Buser & Rappe (1979) studied the pyrolysis of 15 individual
    synthetic PCB congeners and showed that the formation of PCDFs can
    follow several competing reaction pathways. In another study where a
    series of chlorobenzenes were pyrolyzed in the same way, Buser (1979)
    found that significant amounts (> 1%) of PCDDs and PCDFs were formed.
    A complex mixture of isomers of PCDDs and PCDFs was found, suggesting
    several reaction routes. Using the same technique as above, Lindahl et
    al. (1980) studied the thermal decomposition of polychlorinated
    diphenyl ethers. Both PCDDs and PCDFs were formed, involving several
    pathways. The temperature range was 500-600 C and the yields varied
    from 0.1 to 4.5%.

         Bergman et al. (1984) studied the thermal degradation of two
    polychlorinated alkanes containing 59% and 70% chlorine, respectively,
    and also a commercial chlorinated paraffin containing 70% chlorine.
    Their studies indicated the presence of at least mono- and diCDFs.

         Ahling et al. (1978) reported that chlorinated benzenes can be
    found in the pyrolysis of PVC.

         Direct evidence for the conversion of PVC to PCDDs and PCDFs has
    recently been reported by Marklund et al. (1986). They found that
    laboratory pyrolysis of PVC resulted in the formation of PCDDs and
    PCDFs, mainly hexa- and heptaCDDs, and tetra- to heptaCDFs. In some
    cases, the pattern of isomers seemed to be similar to those found in
    municipal and hazardous waste incinerators, e.g. the pentaCDFs (Rappe
    et al., 1987).

         The data discussed in this section are summarized in Table 16.

    3.5.2  Incineration of municipal waste

         For some time, emissions from municipal incinerators, heating
    facilities, and thermal power plants have been the subject of concern.
    Whereas previously the emission of dust, smoke, toxic metals, and
    noxious gases were of prime concern, the presence of potentially
    hazardous organic compounds from these emissions has been recognized
    only recently. Lahaniatis et al. (1977) reported the presence of
    chlorinated organic compounds (chlorinated aliphatics, benzenes, PCBs,
    and pesticides) in fly ash from a municipal incinerator.

         Olie et al. (1977) reported the occurrence of PCDDs and PCDFs in
    fly ash from three municipal incinerators in the Netherlands. Their
    results indicated the presence of up to 17 PCDD peaks, but isomer
    identification and quantification was not possible due to the lack of
    synthetic standards. Buser & Bosshardt (1978) studied fly ash from a
    municipal incinerator and an industrial heating facility, both in
    Switzerland. In the former, the level of PCDDs was 0.2 g/g and of
    PCDFs 0.1 g/g. In the industrial incinerator, the levels were 0.6
    g/g and 0.3 g/g, respectively.

         During the period 1978-1982 a series of papers, reports, and
    reviews were published confirming the original findings of Olie et al.
    (1977) and Buser & Bosshardt (1978) regarding fly ash. Less data have
    been published on the levels of PCDDs and PCDFs in other incineration
    by-products, e.g., particulates and flue gas condensate, and in total
    flue gas, which are the true emissions (Marklund et al., 1986).

         A risk evaluation should be based on the emission levels of PCDD
    and PCDF isomers found in isomer-specific analyses using validated
    sampling and clean-up methods. However, in many studies non-validated
    sampling and analytical methods are used and the results are given in

    terms of total levels of tetra-, penta-, hexa-, hepta-, and octaCDDs
    and CDFs. The value of such studies is limited, particularly in this
    situation where the number of isomers is quite large. More than 30
    PCDDs and 60 PCDFs have been found in fly ash samples (Buser et al.,
    1978b, 1978c).

         In March 1986, a working group of experts convened by the World
    Health Organization Regional Office for Europe reviewed the available
    data on emissions of PCDDs and PCDFs from municipal solid-waste (MSW)
    incinerators. It was found that the origin of these compounds was not
    completely understood, but they appear to result from complex thermal
    reactions occurring during periods of poor combustion. Because of
    their high thermal stability, the PCDDs and PCDFs can be destroyed
    only after adequate residence times at temperatures above 800 C
    (WHO/EURO, 1987).

         Available data on total emissions of PCDDs and PCDFs from tests
    on MSW incinerators range between a few and several thousand ng/Nm3
    dry gas at 10% carbon dioxide (CO2). The working group prepared a
    table giving a range of estimated isomer specific emissions for those
    isomers of major concern with respect to MSW incinerators operating
    under various conditions (Table 17).

         The emissions tabulated in column 1 are those which the working
    group considered to be achievable in the most modern, highly
    controlled, and carefully operated plants in use at the present time.
    Such results do not represent what is considered to be achievable by
    the use of acid gas cleaning equipment; use of such equipment should
    result in much lower values (probably at least one order of
    magnitude). The results given in column 1 are not representative of
    emissions that might be expected from such plants during start-up or
    during occasional abnormal conditions. Emission levels listed in
    column 2 were considered by the working group to be indicative of the
    higher limit of emissions from modern MSW incinerators. These plants
    might experience such emissions during start-up or during occasional
    upset conditions. Consequently, the majority of the available
    concentration data falls between columns 1 and 2. Some of the data
    reviewed has shown that the figures in column 2 should not be
    considered an absolute maximum. However, most existing plants, if
    carefully operated, will have PCDD and PCDF emisions in the range
    between columns 1 and 2.

         The highest values for MSW incinerators (column 3) were obtained
    by multiplying the values in column 2 by a factor of 5. Column 3
    includes emission data that were reported to the working group from
    all tests and under all circumstances. Generally, these emission
    levels are associated with irregular or unstable operating conditions,
    high moisture content of the MSW, low combustion or afterburner
    temperatures, less than adequate technologies, etc.

    
    Table 16.  Formation of PCDDs and PCDFs by thermal processes
                                                                               
    Precursor                      Conditions               Products
                                                                               
    2,4,5-T salt                   Pyrolysis                2,3,7,8-tetraCDD
    2,4,5-T (vegetation)           Pyrolysis                No TCDD
                                   Burning                  No TCDD
    Cl-phenate                     Burning                  PCDDsa + PCDFs
    PCBs                           Pyrolysis                PCDFsb
    PCBzc                          Pyrolysis                PCDFs + PCDDsd
    Cl-Diphenyl ethers             Pyrolysis                PCDFs + PCDDs
    Cl-Alkanes (Paraffins)         Pyrolysis                PCDFs
    PVC                            Pyrolysis                PCDDs + PCDFs
                                                                               

    a = PCDDs formed by dimerization and a non-specific dechlorination.
    b = other products: hexa- and pentaCBs.
    c = polychlorinated benzenes.
    d = other products: PCBs, polychlorinated naphthalenes.
    
         The working group was aware of both lower and higher emission
    levels than those included in Table 17. However, it was felt that the
    values included in Table 17 were likely to be representative of
    emissions from current facilities (WHO/EURO, 1987).

         Of special importance is the observation that the emission of
    1,2,3,7,8-pentaCDD normally exceeds the emission of 2,3,7,8-tetraCDD
    by a factor of three to ten.

    3.5.3  Incineration of sewage sludge

         Sludge from municipal waste water treatment plants may be
    incinerated after being dewatered. The WHO working group (see 3.5.2)
    reviewed the available data from municipal sewage sludge (MSS)
    incinerators, and found that PCDD and PCDF emissions from this type of
    plant were generally lower than emissions from MSW incinerators (see
    Table 17, column 4) (WHO, 1986).

    3.5.4  Incineration of hospital waste

         Doyle et al. (1985) claimed that the incomplete combustion of
    certain hospital waste containing halogenated organics could produce
    high levels of PCDDs and PCDFs. They found the mean values of total
    PCDDs to be 69 ng/m3 and total PCDFs to be 156 ng/m3. No
    isomer-specific data seems to be available. Hagenmaier et al. (1986)
    reported the analyses of stack gas from 10 hospital waste incineration
    plants. The mean value of 2,3,7,8-tetraCCD emitted was 0.28 ng/m3,
    the mean of all TCDDs being 20 ng/m3. The mean value for total PCDDs
    was 118 ng/m3 and for total PCDFs 434 ng/m3.



    
    Table 17. Estimated range of emissions from municipal solid waste (MSW) and municipal sewage sludge (MSS) incineratorsa

                                                                                                                             

                                                               Emissions from MSW combustion                  Emissions
                                                                                                              from MSS
                                                                                                              combustion
                                                                                                                         

                                                            1                2                3                    4
             Congeners                                Achievable with     Maximum            High                 Most
                                                       modern plants       from           emissions              likely
                                                       with no acid       average                                highest
                                                       gas cleaning      operation                              emissions
                                                                                                                         

                                                                 (ng/Nm3, dry, at 10% CO2)
                                                                                                                             

            2,3,7,8-TetraCDD                               0.1              1.5               7.5                 0.1
            1,2,3,7,8-PentaCDD                             0.3             14                70                   0.3
            1,2,3,4,7,8-HexaCDD                            0.2             31               155                   0.2
            1,2,3,6,7,8-HexaCDD                            0.6             56               280                   0.6
            1,2,3,7,8,9-HexaCDD                            0.4             20               100                   0.4
            2,3,7,8-TetraCDF                               0.9             10                50                   0.9
            1,2,3,7,8-/1,2,3,4,8-PentaCDF                  2.3             52               260                   2.3
            2,3,4,7,8-PentaCDF                             2.0             40               200                   2.0
            1,2,3,4,7,8/1,2,3,4,7,9-HexaCDF                1.1             48               240                   1.1
            1,2,3,6,7,8-HexaCDF                            1.3             40               200                   1.3
            1,2,3,7,8,9-HexaCDF                            0.06            52               260                   0.06
            2,3,4,6,7,8-HexaCDF                            2.0             36               180                   2.0
                                                                                                                             

    a     From: WHO/EURO (1987).
    


    3.5.5  Incineration of hazardous waste

         Analyses from a test burn of pentachlorophenol waste have been
    reported by Rappe et al. (1983c). PCP is a well known precursor to
    octaCDD (section 2). Samples of baghouse ash and bottom ash were
    analyzed. In the baghouse ash the total level of octaCDD was only 0.2
    g/g. The major constituents were lower chlorinated PCDDs such as
    hepta-, hexa-, penta-, and tetraCDDs. The isomeric distribution was
    reported to be very similar to a "normal" fly ash. In both cases
    2,3,7,8-tetraCDD was a very minor constituent. The level of PCDD in
    the bottom ash was 0.31 g/g. The baghouse ash was also reported to
    contain PCDFs at a total level of 2.5 g/g. For the tetra- and penta-
    chlorinated compounds, equal amounts of PCDDs and PCDFs were reported.

         Oberg & Bergstrm (1986) reported on test data from a Swedish
    hazardous waste incinerator equipped with a rotary kiln, an
    afterburner, and a dry scrubbing unit. Combustion tests were performed
    with PCB (Aroclor 1242) as a fluid, and as a contaminant in solid
    waste (Aroclor 1016 in capacitors). The results of these tests
    indicated no correlation between the amount of PCB incinerated and the
    amount of PCDDs and PCDFs found in the emissions.

    3.5.6  Metal industry and metal treatment industry

         It has been reported by Marklund et al. (1986) that industrial
    high-temperature processes like copper smelters and electric arc
    furnaces in steel mills have been identified as sources of
    environmental contamination by PCDDs and PCDFs. The results are
    reported in "TCDD equivalents" according to US EPA (1987). The
    emission from the copper smelter contained 11 ng of TCDD
    equivalents/Nm3 dry gas and 10% CO2, while the dust from the steel
    mill contained 0.8 ng TCDD equivalents/g dust. Marklund et al. (1986)
    also considered the emissions from industrial incinerators to be of
    the same magnitude, or even higher, than the emissions from MSW
    incinerators.

         Southerland et al. (1987) analyzed emissions from various
    incinerators within Tier 4 of the USA. The highest levels were found
    in a secondary copper smelter, which contained 170 ng of
    2,3,7,8-tetraCDD/Nm3 and 3% oxygen. This was by far the highest
    level found within the US EPA National Dioxin Strategy.

    3.5.7  Wire reclamation

         Hryhorczuk et al. (1981) studied a wire reclamation incinerator
    in the USA. Using a non-isomer-specific analytical method, they
    determined total levels of tetraCDDs and tetraCDFs. Two samples were
    analyzed, one from the furnace and one from the stack. The furnace
    sample contained 58 ng/kg of total TCDDs and 730 ng/kg of total TCDFs,
    whereas the stack sample contained 410 ng/kg of total TCDDs and 11 600
    ng/kg of total TCDFs.

    3.5.8  Traffic

         Marklund et al. (1987) reported a study where automobile exhaust
    emissions were analyzed for PCDDs and PCDFs. Two groups of test cars
    were utilized: (1) cars equipped with a catalytic converter using
    unleaded gasoline with no halogenated scavengers; (2) cars with no
    catalytic converter using leaded gasoline (0.15 g/litre) and a
    dichloroethane scavenger (0.1 g/litre). Before the test runs, the
    motor oil was changed in all cars. No PCDDs and PCDFs could be
    identifed in the cars using the unleaded gasoline, while the average
    emission from the cars running on leaded gasoline was found to be
    30-540 pg/kg of TCDD equivalents. It was assumed that the chlorinated
    scavenger (dichloroethane) was the precursor of the PCDDs and PCDFs
    formed. It was estimated that the total amount of PCDDs and PCDFs from
    cars in Sweden using leaded gasoline with halogenated scavengers is in
    the range of 10-100 g TCDD equivalents/year.

    3.5.9  Fires and accidents in PCB-filled electrical equipment

         In February 1981 a fire in the State Office Building in
    Binghamton, New York, USA, caused a transformer to rupture, releasing
    soot throughout the building. The dielectric fluid in the transformer
    consisted of a mixture of PCB (65%) and chlorinated benzenes (35%).
    The soot was found to be highly contaminated with PCDFs (total PCDFs
    > 2000 g/g). The most toxic isomers (2,3,7,8-tetraCDF; 1,2,3,7,8-
    and 2,3,4,7,8-pentaCDF; and 1,2,3,4,7,8- and 1,2,3,6,7,8-hexaCDF) were
    found to be the major constituents within each group of congeners.
    Levels reported were 12 mg/g of 2,3,7,8-tetra CDF, 670 mg/g of total
    penta-CDFs, and 965 mg/g of total hexa-CDFs, 46 mg/g of total
    hepta-CDFs, and 460 mg/g of octa-CDFs (Rappe, 1984; Rappe et al.,
    1985b). In addition, a series of PCDDs were identified, including the
    highly toxic 2,3,7,8-tetraCDD, and 1,2,3,7,8-pentaCDD (Rappe et al.,
    1983a; Buser & Rappe, 1984). It is assumed that the chlorinated
    benzenes were the dioxin precursors.

         Between 1981 and 1985, a series of transformer accidents (7 in
    all) similar to the one in Binghamton were reported in the USA and
    Canada (Rappe et al., 1986a). In January 1985, an explosion followed
    by a fire ruptured a transformer in the basement of a residential
    complex in Rheims, France. The transformer was filled with PCB (60%)
    and trichlorobenzene (40%). Total levels of PCDFs were as high as 2570
    g/m2 before clean-up. Only traces of hepta- and octaCDD were found
    (Rappe et al., 1985a).

         In Europe, between 1981 and 1985, 19 accidents involving indoor
    capacitor fires and explosions were reported from Scandinavian
    countries (Rappe et al., 1986a). All capacitors were mineral-oil
    filled, and contamination of the sites averaged 1-5 g total
    PCDFs/m2.

    3.5.10  Pulp and paper industry

         Large amounts of chlorine or chlorine compounds are used in the
    pulp and paper industry for the bleaching of the pulp. Three black
    liquor boilers from the craft paper process were included in the US
    EPA study of combustion sources. No 2,3,7,8-tetraCDD was found in
    these emissions, but low levels of other PCDDs and PCDFs were found in
    one of the three incinerators and a yearly emission of 0.25 g was
    calculated (Southerland et al., 1987).

         Rappe et al. (1987) recently identified both 2,3,7,8-tetraCDD
    (170 ng/kg) and 2,3,7,8-tetraCDF (890 ng/kg) in a sample taken in a
    sedimentation lagoon at a Swedish paper mill. A series of other PCDDs
    and PCDFs was also identified but at lower levels. The isomeric
    pattern in this sample differed markedly from other sediment samples,
    indicating pulping processes to be a source of environmental pollution
    by 2,3,7,8-tetraCDD and 2,3,7,8-tetraCDF (Rappe et al., 1987).

    3.5.11  Incineration of coal, peat, and wood

         The emissions of PCDDs and PCDFs from coal-fired power plants
    (Kimble & Gross 1980), wood stoves (Clement et al., 1985), and peat
    burning (Marklund et al., 1986) seem to be very low when calculated
    per m3. However, the very high flow rates and the large number of
    units could make a significant total contribution. The occurrence of
    pentaCDDs and all PCDFs was not discussed in this report.

    3.5.12  Inorganic chlorine precursors

         It is well known that certain organochlorine compounds are
    efficient precursors to PCDDs and PCDFs during pyrolysis. However, it
    was proposed by scientists from Dow Chemical Company that PCDDs, and
    especially 2,3,7,8-tetraCDD, are ubiquitous and formed as trace level
    by-products of any normal combustion (Bumb et al., 1980).
    Consequently, dioxins should have been present in the environment
    since the advent of fire. This suggests that inorganic chloride can
    serve as a useful precursor to the formation of PCDDs and PCDFs. A
    recent survey of PCDD levels, in particular from residential wood
    combustion units, has been quoted in support of the above. The survey
    showed PCDD levels in the ng/kg range (see also section 3.5.11).
    However, this hypothesis has been criticized. One of the main
    arguments against such a hypothesis is that 2,3,7,8-tetraCDD does not
    appear to be formed in coal-fired power plants (Kimble & Gross, 1980;
    Junk & Richard, 1981). Another argument is that the Dow studies lack
    data on levels of dioxin precursors in the material being burned,
    including the air in the flames (Rappe, 1984).

         The analyses of historical samples gives additional support to
    the theory that organochlorine compounds are more important as
    precursors than inorganic chloride. When Czuczwa & Hites (1985)
    analyzed sediment core samples from Lake Huron, N. America, the first
    indication of PCDDs and PCDFs was found in sediments from 1940. There
    was also a good correlation between the trend in the levels of PCDDs
    and PCDFs in these sediments and the trend in the production of
    chlorinated aromatic compounds (section 5.4).

         Schecter et al. (1986a) were unable to detect PCDDs and PCDFs in
    human liver and lung tissue from two female Eskimos frozen over one
    hundred years ago (see also section 4.4.4).

    3.5.13  Photochemical processes

         Sundstrm et al. (1979) studied the formation of 2,3,7,8-tetraCDD
    in six re-forestation areas that were sprayed with 2,4,5-T esters.
    Leaf samples from the areas were analyzed for 2,4,5-T esters and TCDD.
    TCDD was found in one leaf sample only, at levels lower than expected
    from the level of dioxin contamination of the herbicide formulations
    used.

         The photochemical formation of PCDDs and PCDFs has also been
    studied in laboratory experiments.

         The photochemical dimerization of chlorophenols to PCDDs was
    studied by Crosby & Wong (1976). The only PCDD formed in this study
    was the octaCDD. Other PCDDs can be formed by photochemical
    cyclization of chlorinated o-phenoxyphenols, also called pre-dioxins
    (Nilsson et al., 1974). These pre-dioxins are very common impurities
    (1-5%) in commercial chlorophenols (Nilsson et al., 1978), but the
    cyclization is only a minor reaction pathway; the main reaction being
    the dechlorination of the pre-dioxin.

         Akermark (1978) studied the formation of 2,3,7,8-tetraCDD from
    the appropriate pre-dioxins. He could identify the product, but
    claimed the reaction to be very inefficient.

         Another photochemical process of potential environmental
    importance is dechlorination of the higher chlorinated PCDDs and
    PCDFs, e.g., octaCDD and octaCDF. The products formed by photolysis of
    octaCDD in organic solvent have now been identified (Buser & Rappe,
    1978). By comparison with authentic standards, it was found that the
    main tetrachloro isomer was the 1,4,6,9-tetraCDD; the major
    pentachloro compound was expected to be the 1,2,4,6,9-isomer, and the
    main hexa- and heptachloro compounds were the 1,2,4,6,7,9- (or
    1,2,4,6,8,9-) and the 1,2,3,4,6,7,9-isomers, respectively. The
    reaction scheme deduced from this data indicates that the chlorine

    atoms are removed preferentially from the lateral positions on the
    carbon rings. Consequently, the most toxic PCDD isomers, such as
    2,3,7,8-tetraCDD, are not likely to be formed from the photolysis of
    the higher PCDDs in solution.

         Crosby et al. (1973) studied the photolysis of a series of PCBs
    dispersed in water. For two isomers, the 2,5-dichloro- and
    2,2',5,5'-tetrachlorobiphenyls, small amounts (0.2%) of 2-mono-CDF
    could be found among the products (for photo-chemical transformations,
    see section 4.2.1).

    3.6  Comparison of Isomeric Pattern and Congener Profiles From Various Sources

         There is a pronounced difference between technical products and
    incineration emissions in both isomeric patterns and congener profiles
    of PCDDs and PCDFs. In technical products the number of isomers
    present is limited, whereas in incineration emissions most isomers
    seem to be present. Rappe (1987) has pointed out the large similarity
    qualitatively in isomeric patterns between different incineration
    sources.

    4.  ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATIONS

    4.1  Environmental Transport

    4.1.1  Air

         The PCDDs and PCDFs are believed to be transported in the
    atmosphere. The transport of these compounds from stacks and other
    stationary point sources, as well as from waste disposal sites and
    other area sources, can be predicted from dispersion modelling (SAI,
    1980). In the case of the accidental release of a toxic cloud
    containing 2,3,7,8-tetraCDD at Seveso, Italy, Cavallaro et al. (1982)
    determined the transport pattern and the ground deposition. They
    determined that the deposition of 2,3,7-8-tetraCDD from air to soil
    should follow an exponential decay pattern in the
    Gaussian-distribution along the cross-section of the downwind
    direction. Thibodeaux (1983) studied the air transport of
    2,3,7,8-tetraCDD at a herbicide production facility in Jacksonville,
    Arkansas, USA.

         The dispersion modelling has limitations. If possible, the
    modelling calculations should be combined with true air measurements.

    4.1.2  Water

         The solubility of 2,3,7,8-tetraCDD in water has been extensively
    studied (see section 2), but much less data are available for the
    other PCDDs and PCDFs. However, data from microbiological experiments
    indicate that 2,3,7,8-tetraCDD is highly adsorbed to sediments and
    biota. Matsumura et al. (1983) suggested that more than 90% of the
    2,3,7,8-tetraCDD in an aquatic medium could be present in the adsorbed
    state. Rappe et al. (1985c) studied a suspension of soot/dust in the
    wash water from a PCB fire. The suspension contained 100 ng/ml of
    various PCDFs, but when the soot was settled the water contained no
    detectable levels of PCDFs (detection level: 0.1 ng/ml of each
    isomer). Most of the PCDDs and PCDFs, if present in waterways, should
    be in the sediments or attached to suspended particles.

         Thibodeaux (1983) has calculated the amount of 2,3,7,8-tetraCDD
    transported by a creek within the contaminated herbicide factory in
    Jacksonville, Arkansas, USA. The value was 0.89 g/year as an average
    rate, and a maximum of 2.1 g/year.

    4.1.3  Soil and sediments

         The mobility of 2,3,7,8-tetraCDD and of a dichlorodioxin in soils
    has been studied by Helling et al. (1973). Both were found to be
    immobile in all soils and, therefore, would not be leached out by
    rainfall or irrigation, though lateral transport during surface
    erosion of the soil could occur.

         The US Air Force conducted studies in an area of north-west
    Florida that had been heavily sprayed with the herbicide Agent Orange
    between 1962 and 1964 (Young et al., 1975). This herbicide mixture was
    contaminated with TCDD (section 3.2). A 7.8-ha test grid received a
    total of 40 metric tons of 2,4,5-T between 1962 and 1964. When 15-cm
    soil core samples were taken in 1974, they showed TCDD concentrations
    ranging from 10 to 710 ng/kg. This study illustrates that significant
    levels of TCDD residues remained 10 years after the last herbicide
    application. Similar TCDD concentrations were obtained from areas that
    had been sprayed between 1962 and 1969 (Bartelson et al., 1975).

         In another study Young (1983) measured the concentration of
    2,3,7,8-tetraCDD in a soil profile. The samples were collected in 1974
    and the data suggested that most of the 2,3,7,8-tetraCDD would be
    found in the top 15 cm of the soil profile (Table 18).

         The probable media and modes of transport of PCDDs from soils are
    the following: (1) to air via contaminated airborne dust particles;
    (2) to surface water via eroded soil transported by water; (3) to
    groundwater via leaching; (4) to air via volatilization. Movement of
    particulate matter containing adsorbed PCDDs and PCDFs has been
    considered to be a much more important transport mechanism than
    leaching and volatilization because of the low water solubility and
    volatility of these compounds (Josephson, 1983). However, the
    monitoring of Seveso soil one year after the accident showed that the
    highest 2,3,7,8-tetraCDD levels were not present in the topmost soil
    layer (0.5 cm), but very often in the second (0.5-1.0 cm) or third
    (1.0-1.5 cm) layers. This disappearance of at least a part of the
    2,3,7,8-tetraCDD from the topmost soil layer was speculated to be due
    to volatilization or vertical movement through the soil (DiDomenico et
    al., 1980). Therefore, it appears that volatilization from soil and
    leaching to groundwater can be responsible for the transport of PCDDs
    and PCDFs from soils under certain conditions, namely, heavy rainfall
    on sandy soils. Studies by Young (1983) indicate that the half-life
    for 2,3,7,8-tetraCDD in soil is 10-12 years.

         Thibodeaux (1983) has calculated the vaporization of
    2,3,7,8-tetraCDD from a herbicide plant in Jacksonville, Arkansas,
    USA. The vaporization can take place from soil surfaces, from landfill
    cells, and from the surface of a pond. In Table 19 a summary of yearly
    emission rates from these sources is presented.

         It was found that vaporization from the soil surface in the
    highly contaminated blow out area was the major contributing source of
    emissions from this plant.

    
    Table 18. Concentration of 2,3,7,8-tetraCDD in a soil profilea, b
                                                                               
               Depth (cm)                        2,3,7,8-tetraCDD (ng/kg)
                                                                               
               0   - 2.5                                    150
               2.5 - 5.0                                    160
               5.0 - 10                                     700
               10  - 15                                      44
               15  - 90                                      NDc
                                                                               

    a     From: Young (1983).
    b     The area received 1,069 kg/ha of 2,4,5-T Agent Orange during
          1962-1964. The soil samples were collected and analyzed in 1974.
    c     None detected (minimum detection limit: 10 ng/kg).
    

        Table 19. Surface source areas and emission rates of 2,3,7,8-tetraCDDa

                                                                               
    Source                               Area (m2)     Emission rate (g/year)
                                                                               

    Blow-out area, volatilization           753                 120-1200
    Blow-out area, entrainment              753                  28-37
    Rocky Branch Creek, dissolved                              0.89-2.1
    Reasor-Hill dump                       1129                 0.1-1.0
    Rocky Branch Creek, sediment                              0.094-0.22
    Cooling water pond                    15050               0.015-0.016
    Total                                                       150-1240

                                                                               

    a From: Thibodeaux (1983).
    
         Freeman et al. (1986) have developed a model to describe the
    vaporization and diffusion through a column of soil of low volatility
    organic chemicals like PCDDs and PCDFs. This model has been used to
    make predictions on the transport of 2,3,7,8-tetraCDD at a site in
    Times Beach, Missouri, USA. The model predicted that the 1983 levels
    in this soil would be only 10% of the original loading. The model also
    predicted that 57% of the initial amount of 2,3,7,8-tetraCDD was
    vaporized through the soil column to the surface in the first year
    after the spraying and that most transport of the vapour occurred
    during the summer months. The same results were also obtained in
    studies reported by Facchetti et al. (1986) and Palausky et al.
    (1986).

    4.2  Environmental Transformation

    4.2.1  Abiotic transformation

         Like other PCDDs and PCDFs, 2,3,7,8-tetraCDD is chemically quite
    stable, and is not likely to be degraded at a significant rate by
    hydrolytic reactions under environmental conditions. Under these
    conditions, TCDD seems also to be rather stable to photochemical
    degradation (Crosby et al., 1971). The half-life of TCDD of about
    10-12 years, as found by Young et al. (1983) for soil, is in agreement
    with this observation.

         However, three reports on rapid photochemical degradation of
    2,3,7,8-tetraCDD under experimental conditions make the situation more
    complicated. In a methanol solution, TCDD is fairly easily degraded by
    photolysis in the laboratory (Crosby et al., 1971). Other studies
    using 2,4,5-T ester formulations with known amounts of TCDD and
    exposed to natural sunlight on leaves, soil, or glass plates showed
    that most of the TCDD was lost during a single day (Crosby & Wong,
    1977). In these two studies, a "hydrogen donor", such as methanol or
    2,4,5-T ester, enhanced the photochemical dechlorination (Akermark,
    1978); they do not therefore truly reflect environmental conditions,
    where the 2,4,5-T ester would be rapidly hydrolyzed on the surface of
    the leaves. At Seveso the TCDD was released together with salts of
    2,4,5-trichlorophenol, ethylene glycol, and inorganic constituents
    (Rappe, 1978b). Like water, none of these is a potent hydrogen donor.

         According to Bertoni et al. (1978), the addition of a solution of
    ethyl oleate in xylene enhances the breakdown of TCDD in soil by UV
    light. Similarly, a cationic surfactant, 1-hexadecylpyridinium
    chloride, was also reported to enhance photodecomposition (Botre et
    al., 1978).

         Another experiment, which might be a good model for the
    degradation of TCDD bound to dust particles in the air, has shown that
    TCDD adsorbed on silica gel undergoes rapid photo-chemical degradation
    (Gebefugi et al., 1977).

         In order to explain the longer half-life of 2,3,7,8-tetraCDD in
    a model laboratory ecosystem than in an outdoor pond, Matsumura et al.
    (1983) speculated that photolysis was the most likely cause. In the
    outdoor environment, algae-mediated photosensitization of
    2,3,7,8-tetra-CDD may have caused some photodecomposition of this
    compound.

         An increase in chlorine substitution is expected to decrease the
    rate of photodegradation. For example, Crosby et al. (1971) showed
    that although complete decompostion of 2,3,7,8-tetraCDD in methanol
    occurred in 24 h under UV irradiation, > 80% octaCDD in methanol
    remained unreacted during the same period under similar irradiation
    conditions.

         Although the degree of photolysis may be related to the extent of
    chlorination, different chlorine substitution patterns also play a
    critical part. In higher chlorinated PCDDs, there appears to be
    preferential loss of chlorine from the 2,3,7, and 8 positions (Buser
    & Rappe, 1978). Thus, PCDDs with chlorine substitutions in positions
    2,3,7, and 8 are likely to be photochemically degraded faster than
    compounds not having these positions substituted. For example, the
    photolysis half-life of 1,2,3,7,8-pentaCDD has been estimated to be
    7.8 h in n-hexadecane solution under sunlamp irradiation (Nestrick
    et al., 1980). Similarly, the photolytic half-lives of
    1,2,3,7,8-pentaCDD, 1,2,3,6,7,9-, and 1,2,4,6,7,9-hexaCDD in hexane
    solutions under sunlight irradiation have been determined to be 5.4,
    17, and 47 hours, respectively (Dobles & Grant, 1979). Nestrick et al.
    (1980) reported a half-life value of 6.8 h for 1,2,3,6,7,8-hexaCDD in
    n-hexadecane under sunlamp irradiation. The primary intermediates of
    the photo-degradation of higher chlorinated PCDDs are probably lower
    chlorinated dioxins (Buser & Rappe, 1978), but the pathways of
    degradation are not known with certainty (National Research Council of
    Canada, 1981).

         From these discussions of the photolysis of PCDDs in the presence
    of organic hydrogen-donating substrates, it is difficult to predict
    the photolytic fate of these compounds in natural aquatic media, where
    hydrogen donors may or may not be available. The situation is
    complicated further by the fact that a predominant amount of PCDDs in
    surface water may be adsorbed or suspended on particles and sediments,
    rather than in solution. Moreover, since the penetration of UV light
    into natural water may be very limited, photolytic degradation of
    PCDDs in water is not likely to be of environmental importance.

         Hutzinger et al. (1973) have studied the photochemical
    degradation of 2,8-diCDF and octaCDF. They found that a reductive
    dechlorination takes place, especially in methanolic solution. The
    reaction was much slower when a thin film was exposed to sunlight.

         Thermally, 2,3,7,8-tetraCDD is quite stable, rapid decomposition
    occurring only at temperatures above 750 C (Stehl et al., 1973).

    4.2.2  Biotransformation and biodegradation

         The 2,3,7,8-tetraCDD isomer is very resistant to biodegradation.
    Only 5 of about 100 microbial strains with the ability to degrade
    persistent pesticides were able to degrade 2,3,7,8-tetraCDD (Matsumura
    & Benezet, 1973). Ward & Matsumura (1978) studied the biodegradation
    of 14C-labelled 2,3,7,8-tetraCDD in lake waters and sediments from
    Wisconsin, USA, and observed a half-life of 2,3,7,8-tetraCDD in lake
    waters containing sediment of 550-590 days. In lake water alone, about
    70% of the 2,3,7,8-tetraCDD remained after 589 days. Using an outdoor
    pond as a model aquatic ecosystem, and dosing it with 14C-labelled
    2,3,7,8-tetraCDD, Matsumura et al. (1983) estimated the half-life of
    2,3,7,8-tetraCDD to be approximately 1 year. Although biodegradation

    may have been responsible for part of the degradation, it is almost
    impossible to estimate the biodegradation half-life of
    2,3,7,8-tetraCDD in aquatic systems from this experiment.

         Philippi et al. (1982) detected a polar metabolite of
    2,3,7,8-tetraCDD in several microbiological cultures after long-term
    incubation. They reported chromatographic and MS data that supported
    the conclusion that the metabolite was 1-hydroxy-2,3,7,8-tetraCDD,
    although a synthetic standard compound was not available.

         Tulp & Hutzinger (1978) reported that in rats, dibenzo-p-dioxin,
    1-monoCDD, 2-monoCDD, 2,3-diCDD, 2,7-diCDD, 1,2,4-triCDD, and
    1,2,3,4-tetraCDD are metabolized to mono- and di-hydroxy derivatives.
    In the case of dibenzo-p-dioxin and both of the two monochloro 
    isomers, sulfur-containing metabolites were also excreted. Primary
    hydroxylation exclusively took place at the 2, 3, 7, or 8 positions in
    the molecule. In these studies, no metabolites resulting from fission
    of the C-O bonds (ortho, ortho'-dihydroxychlorodiphenyl ethers,
    chloro-catechols), or hydroxylated derivatives thereof, were detected.
    No metabolites were found from octaCDD.

    4.3  Bioaccumulation

         The bioaccumulation of 2,3,7,8-tetraCDD has been investigated in
    several studies, using several aquatic species and different model
    ecosystems. In the experiments in which 14C-TCDD was introduced into
    the model ecosystem in the form of residues on sand, particularly high
    values were found in the mosquito (Aedes egypti) larvae, the level
    exceeding that found in water by more than 9000 times. Under similar
    conditions, the level in brine shrimp (Artemia salina) was 1570 times
    higher than that found in water (Matsumura & Benezet, 1973). In the
    second study (Isensee & Jones, 1975; Isensee, 1978), 14C-TCDD was
    absorbed, at a broad range of levels, into soil and placed at the
    bottom of an aquarium. Five species of organisms were added (though
    not simultaneously) 1-30 days after flooding, and exposed for 3-32
    days. The correlation between the TCDD level in the water and in the
    organisms of each species was highly significant (correlation
    coefficient of 0.94 or higher).

         Bioaccumulation factors for 2,3,7,8-tetraCDD are given in Table
    20 (US EPA, 1985).

    4.4  Levels in Biota

    4.4.1  Vegetation

         When 14C-labelled 2,3,7,8-tetraCDD was added to soil, both oats
    and soya beans accumulated small quantities of TCDD, at all stages of
    growth. TCDD was also detected in control plants housed with the
    experimental plants after treatment (Isensee & Jones, 1971). A maximum
    of 0.15% of the TCDD present in the soil was translocated to the

    aerial portion of the oats and the soya beans, but neither the grain
    nor the beans harvested at maturity showed any detectable level of
    14C-labelled TCDD. When TCDD was applied to the central leaflet of
    3-week-old soya bean plants and 12-day-old oat plants, very little
    TCDD was lost from the soya bean leaves in 21 days, but there was a
    gradual loss (38% in 21 days) from the oat leaves.

         Analyses of vegetation from Seveso, Italy, after the industrial
    accident, gave values of up to 50 mg TCDD/kg (Firestone, 1978). In the
    following years, when there was no direct contact of the newly grown
    vegetation with the aerosol cloud, the levels of dioxin in plants
    decreased by several orders of magnitude (Wipf & Schmid, 1983). In
    1977 (one year after the accident in Seveso), no traces of TCDD were
    found in the flesh of apples, pears, and peaches, or in corn cobs or
    kernels, grown near the factory (the detection limit for the analyses
    was 1 ng/kg). At the same time about 100 ng TCDD/kg was detected in
    the fruit peels. This strongly suggests that the contamination was due
    to dust and not from plant uptake. The TCDD level in the soil was
    found to be in the order of 10 ng/g, which corresponds to about 1000
    g/m2 (Wipf et al., 1982).

         Facchetti et al. (1986) studied plants grown in soil spiked with
    2,3,7,8-tetraCDD in the range 1-752 ng TCDD/kg. At the end of
    cultivation, root samples were collected, carefully washed, and
    analyzed. The levels of TCDD in the roots were found to be higher than
    the levels found in the soil in which the plants were grown. On the
    parts above ground, Facchetti et al. (1986) could not find any
    significant increase in the levels of TCDD. However, the TCDD
    concentration was found to vary with the location, being higher if the
    plants were grown in the vicinity of other pots containing
    contaminated soil. The conclusion was drawn that evaporation is the
    predominant process for the contamination of the aerial parts of
    plants. However, studies by Sacchi et al. (1986) indicated that maize
    and bean plants grown in soil contaminated by 3H-2,3,7,8-tetraCDD
    accumulated radioactivity in the aerial parts progressively with time
    and with soil contamination (Sacci et al., 1986). It was suggested
    that the distribution of the TCDD into the leaves occurred via the
    transpiration stream.

         Very few analyses of sprayed vegetation have been reported. A
    rough estimate of 20-1000 ng/kg for 2,3,7,8-tetraCDD contamination can
    be made on the basis of the level of 2,4,5-T found in newly sprayed
    vegetation and the level of 2,3,7,8-tetraCDD in the spray formulation
    used. Higher values could be obtained for Agent Orange. Sundstrm et
    al. (1979) reported data in agreement with this estimate. However, the
    analytical technique used in their study was not isomer specific.
    Vegetation was sprayed with 2,4,5-T ester contaminated by only 0.06 mg
    2,3,7,8-tetraCDD/g. A sample of leaves collected 42-45 days after the
    spraying was found to have 170 ng TCDD/kg, somewhat lower than the
    expected value 600 ng TCDD/kg, indicating a slow photochemical
    breakdown.



    
    Table 20. Measured bioaccumulation factor for 2,3,7,8-TCDD in freshwater aquatic organismsa
                                                                                                                             

          Species                       Tissue         Duration       Bioconcentration                 Reference
                                                        (days)             factor
                                                                                                                             
          Alga                                            33               3094b                       Isensee (1978)
          (Oedogonium cardiacum)

          Alga                                            32               2075c                       Isensee (1978)
          (Oedogonium cardiacum)                                                                       Yockim et al. (1978)

          Snail                       whole body          33               5471b                       Isensee (1978)
          (Physa sp.)

          Snail                       whole body          32               3095c                       Isensee (1978)
          (Physa sp.)                                                      3731                        Yockim et al. (1978)

          Cladoceran                  whole body          32               3895b                       Isensee (1978)
          (Daphnia magna)

          Cladoceran                  whole body          30               7070c                       Isensee (1978)
          (Daphnia magna)                                                  7125                        Yockim et al. (1978)

          Catfish                     whole body          28               4875                        Yockim et al. (1978)
          (Italurus punctatus)

          Mosquitofish                whole body          14               4850c                       Isensee (1978)
          (Gambusia affinis)                                               4875                        Yockim et al. (1978)



    a     From: US EPA (1985).
    b     Arithmetic mean of several values reported.
    c     Tissue concentrations at equilibrium.
    


    
    Table 21.  Levels of TCDDs in fish and shellfisha

                                                                               
         Sample            Tissue                    Concentration of
         number            typeb                     2,3,7,8-TCDD (ng/kg)c
                                                                               

           1          Fish (edible flesh)                480
           2          Catfish                             40
           3          Buffalo fish                        ND(13)
           4          Fish (predator)                    230
           5          Fish (bottom feeder)                77
           6          Catfish                             50
           7          Buffalo fish                        ND(7)
           8          Catfish                             ND(7)
                                                                               

    a     From: Mitchum et al. (1980).
    b     All samples were obtained from the Arkansas River, USA, or
          a tributary, the Bayou Meto.
    c     These are averages of samples that had detectable levels of
          TCDD.
    ND   = none detected; the number in parenthesis is the measured
           detection limited for that sample.
    

    4.4.2  Aquatic organisms

         Fish and shellfish taken from areas in South Viet Nam that were
    heavily exposed to Agent Orange during military defoliation operations
    in the 1960s have been reported to contain 18-810 ng TCDD/kg (Baughman
    & Meselson, 1973). The analytical technique of direct-inlet high
    resolution MS used in this study is not considered isomer specific and
    did not include any GC separation at all.

         In two streams associated with the US Air Force test area in
    north-west Florida (section 4.1.3), which had been heavily sprayed
    with Agent Orange between 1962 and 1964, the silt contained, 10 years
    later, 10 and 35 ng TCDD/kg where eroded soil entered the water.
    Concentrations of 12 ng TCDD/kg were found in two species of fish from
    this stream, the sailfin shiner (Notropis hypselopterus) and the
    mosquito fish (Gambusia affinis). The spotted sunfish (Lepomis
    punctatus) contained 4 ng TCDD/kg in skin and muscle, 18 ng/kg in the
    gonads, and 85 ng/kg in the gut (Young et al., 1976).

    
    Table 22. Analytical results for 2,3,7,8-tetraCDD residues in fish from Saginaw Bay Region, Michigan, USAa

                                                                               

    Species    Number of       Number of              TCDD
               samplesb        positive             detected (ng/kg)c
                               samples              low   high   mean
                                                                               
    Channel
    catfish     8               8                    28    695    157 (13)

    Carp       14              10                    20    153     55 (7)

    Yellow
    perch       6               3                    10     20     13 (5)

    Smallmouth
    bass        2               2                     7      8      8 (6)

    Sucker      4               3                     4     21     10 (4)

    Lake trout  2               0                     0      0      0 (5)
                                                                               

    a     From: Harless et al. 1982).
    b     Mean % recovery for 2.5-10 ng 37Cl4-TCDD added to 5 or 10
          g of tissue prior to sample preparation was between 78 and
          100%.
    c     Corrected for losses in efficiency of sample preparation for
          particular species. The numbers in parenthesis indicate the
          limit of detection for TCDD.
    
         The levels of TCDD in fish from the Atlantic or from ponds in the
    USA in areas sprayed with 2,4,5-T were below the detection levels (1-2
    ng/kg) (Baughman, 1974; Shadoff et al., 1977).

         Mitchum et al. (1980) reported levels of 400 ng
    2,3,7,8-tetraCDD/kg in fish samples collected in Bayou Meto/Arkansas
    River, USA, a waterway associated with industrial plants for the
    production of 2,4,5-T (Thibodeux, 1983) (see Table 21).

         Levels ranging from 4-695 ng TCDD/kg were found in the edible
    portion of channel catfish, carp, yellow perch, small-mouth bass, and
    suckers from Saginaw Bay, Michigan, USA, near facilities used for the
    production of 2,4,5-T herbicides. The highest concentrations were
    detected in bottom-feeding catfish and carp, while the lowest
    concentrations were detected in bass, perch, and suckers (see Table
    22) (Harless et al., 1982).

         Rappe et al. (1981) identified a series of tetra- to octaCDFs in
    fat samples of a snapping turtle from the Hudson River and of gray
    seal from the Baltic Sea. The total levels of PCDFs in these samples
    were 3 ng/g and 40 ng/kg, respectively. In both samples the major
    PCDFs consisted of the most toxic isomers (2,3,7,8-tetra-;
    2,3,4,7,8-penta-; and 1,2,3,4,7,8- and 1,2,3,6,7,8-hexaCDFs).

         Norstrom et al. (1982) have analyzed pooled samples of herring
    gull eggs collected in 1982 from various parts of the Great Lakes, N.
    America. In all samples, 2,3,7,8-tetraCDD was found in levels ranging
    from 9 to 90 ng/kg. The identity of the 2,3,7,8-isomer was confirmed
    by retention times on three capillary columns. In another study
    Stalling et al. (1983) were not able to detect measurable levels of
    tetraCDDs and other PCDDs in fish samples from Lake Superior,
    N.America (the detection level was 2-5 pg/g). The difference could be
    explained by the migration of the herring gulls during the winter. On
    the other hand, a series of PCDFs could be identified in the Lake
    Superior fish samples, indicating more widespread background levels
    for the PCDFs than for the PCDDs. Stalling et al. (1983) found the
    total levels of PCDFs in fish samples from Lakes Michigan, Huron, and
    Ontario, N. America, to be 12-290 ng/kg. The toxic 2,3,7,8-substituted
    PCDDs and PCDFs were present in all samples, the highest levels being
    found in samples from Lake Huron, Lake Ontario, and the Tittabawasee
    River, which flows into Saginaw Bay. The residue pattern found in the
    fish and locally high levels suggest a strong influence by local point
    source discharges (Stalling et al., 1983). The data of O'Keefe et al.
    (1983) are also in agreement with this theory.

         Norstrom et al. (1986) have studied the long-term trends of
    2,3,7,8-substituted PCDDs and PCDFs in herring gull eggs in the Great
    Lakes. The levels of 2,3,7,8-tetraCDD were found to decline
    exponentially in Lake Ontario, with a half-life of 3-4 years, from a
    high of 2000-5000 ng/kg in the early 1970s to a level of 80-100 ng/kg
    in 1984/1985. The levels of TCDD in Lake Michigan were 249 ng/kg in
    1971, 70 ng/kg in 1972, and 10-20 ng/kg in 1984/1985. These levels
    have not changed significantly since 1979. This suggests that an
    equilibrium between input and removal mechanisms has been established
    in this water system for most PCDDs and PCDFs. The same trend is
    reported for various fish species in Lake Ontario (Ontario, 1986).

         Ryan et al. (1983a) analyzed a series of commercial and sport
    fish from the Great Lakes and from the Pacific coast of Canada for
    2,3,7,8-tetraCDD (Table 23). The highest levels were found in Lake
    Huron and Lake Ontario. In a preliminary study they also reported
    finding levels of 2,3,7,8-tetraCDFs and other unidentified tetraCDFs
    of 3-200 ng/kg of fish.

         The Baltic Sea is an area of interest because this region is
    without any known point sources of dioxins. Rappe et al. (1987)
    reported on the analyses of two samples of homing salmon and two
    samples of pooled herring; one herring sample from the Baltic Sea
    (Karlskrona) and the other from the northern part of the Gulf of
    Bothnia (Lulea) (Table 24). As expected, the levels in the salmon
    muscle were much higher than the levels found in the herrings, but,
    unexpectedly, the levels in the herring sample from the Gulf of
    Bothnia (Lulea) were somewhat higher than levels found in the sample
    from the Baltic Sea (Karlskrona).

         An interesting observation is that in the majority of the aquatic
    samples only the 2,3,7,8-substituted PCDD and PCDF congeners were
    found. However, crustaceans seemed to be an exception from this
    general trend. Norstrm et al. (1988) reported that crab
    hepatopancreas from the Canadian Pacific Coast contain other
    congeners, e.g. 1,2,4,7,8-pentaCDD and
    1,2,3,6,7,9-/1,2,3,6,8,9-hexaCDD. Rappe et al. (1987) collected and
    analyzed crab hepatopancreas from three different locations along the
    west coast of Sweden. The crab samples from the locations Grebbestad
    and Idefjord should represent background levels, while Vrfjord has
    a potential point source of dioxins from a pulp mill using chlorine
    for bleaching. The results are given in Table 25.

         Low background levels of series PCDDs and PCDFs were found in all
    samples. In addition, the sample from the Vrfjord also contained
    much higher levels of some congeners, especially 2,3,7,8-tetraCDF and
    2,3,7,8-tetraCDD. This is another indication that pulp bleaching could
    be a potential source of 2,3,7,8-tetraCDD and 2,3,7,8-tetraCDF (see
    section 3.5.10).

    4.4.3  Terrestrial animals

         In a heavily sprayed test area in north-west Florida (Young et
    al., 1976), a total of 106 adult and 67 fetuses of beach mice
    (Peromysous polionotus) were collected in 1973 and 1974 and
    examined (method not specified). Livers from the beach mice contained
    from 540-1300 ng TCDD/kg and the pelts 130-140 ng/kg. The visceral
    mass of race runners (Cnemidophorus sexlineatus) which were caught
    in that area contained 360 ng TCDD/kg and the trunk of the reptiles
    contained 370 ng/kg.

         At the time of the accident in Seveso, Italy, more than 81 000
    animals were inhabiting the contaminated zones. Most were rabbits (25
    000), poultry, and other small animals (55 500), with 349 cattle, 233
    pigs, 49 horses, 21 sheep, and 49 goats also in the zones. Many of
    these animals died and others were killed. A large number of these
    animals were analyzed for 2,3,7,8-tetraCDD by a method with a
    detection level of 250 ng/kg (Pocchiari et al., 1983). The results are
    summarized in Tables 26 and 27.

        Table 23. Levels of 2,3,7,8-tetraCDD and PCB in Great Lakes Canadian sport
    fish (1980) and smelt (1979)a
                                                                               
    Species             Origin                 TCDD              PCB
                                              (ng/kg)           (g/g)
                                                                               
    Lake troutb         Lake Ontario            58               7.28
                        Lake Huron              37               5.03

    Rainbow troutb      Lake Ontario            33               1.77

    Coho salmon         Lake Ontario            28c              7.39
                        Pacific Coast           NDd (4)          0.03

    Smelt               Lake Ontario            11
                                                16
                                                11
                        Lake Erie               NDd (2)
                                                                               

    a     From: Ryan et al. (1983a).
    b     Whole fish.
    c     Also contained 36 ng hexaCDD/kg (three isomers) and 93 ng
          octaCDD/kg.
    d     ND = not detected at bracketed detection limit.
    
        Table 24. Levels of PCDDs and PCDFs in fish samples from the Baltic Sea
    (pg/g) a,b
                                                                               
                                      Salmon      Salmon      Herring    Herring
                                      Ume River   Ume River   Karlskrona Lulea
                                       1985        1985        1983       1983
                                                                               
    2,3,7,8-TetraCDF                  29          12            5.5       3.0
    2,3,7,8-TetraCDD                   1.9         1.3        < 0.3     < 0.6
    1,2,3,7,8-/1,2,3,4,8-PentaCDF      6.9         3.3          1.4       0.9
    2,3,4,7,8-PentaCDF                49.0        23.0          6.8       8.8
    1,2,3,7,8-PentaCDD                 8.8         4.3          1.1       4.7
    1,2,3,4,7,8-/1,2,3,4,7,9-HexaCDF   1.1         0.7          0.4       0.3
    1,2,3,6,7,8-HexaCDF                1.3         0.8          0.4       0.3
    1,2,3,7,8,9-HexaCDF               ND          ND            0.4       0.2
    2,3,4,6,7,8-HexaCDF                1.1         0.6          0.4       0.2
    1,2,3,4,7,8-HexaCDD               ND           0.4          0.2      ND
    1,2,3,6,7,8-HexaCDD                4.6         2.3         ND         8.1
    1,2,3,7,8,9-HexaCDD               ND          ND           ND        ND
    Total HeptaCDFs                   ND           2.7          0.8      ND
    Total HeptaCDDs                   ND          ND           ND        ND
    OctaCDF                           ND           1.0         ND        ND
    OctaCDD                           ND          ND           ND        ND
                                                                               

    Table 24.(cont'd)   Levels of 2,3,7,8-tetraCDD and PCB in Great
    Lakes Canadian sport fish (1980) and smelt (1979)a
                                                                               

    a     From: Rappe et al. (1987).
    b     ND indicates a level < 0.1 pg/g.
    
         Harless et al. (1983) reported a study in which 2,4,5-T
    containing less than 0.1 mg of 2,3,7,8-tetraCDD/kg was applied at a
    rate of 3.4 kg/ha to approximately 3 ha of an enclosed plot (4.5 ha).
    Twelve deer were placed in the enclosure prior to the application of
    2,4,5-T. One deer died two days later of unknown causes. The remaining
    deer were sacrificed prior to, and at specific intervals during, the
    course of the 30-day study. The analytical results are summarized in
    Table 28.

         In another study (Hryhorczuk et al., 1981), samples from a horse
    grazing close to a wire reclamation incinerator were analyzed and
    found to contain unspecified tetraCDFs (165 ng/kg in the fat, 57 ng/kg
    in the liver) and unspecified tetraCDDs (45 ng/kg in the fat and less
    than 6 ng/kg in the liver) (compare section 3.5.7).

         In order to identify PCDD and PCDF levels in the general
    terrestrial background, Nygren et al. (1986) analyzed bovine samples
    - fat, liver, and milk - and identified the same 2,3,7,8-substituted
    PCDDs and PCDFs as were found in the aquatic samples. However, the
    levels were lower and normally close to the detection limit.

    4.4.4  Human data

         Occupational exposure to 2,3,7,8-tetraCDD can occur during the
    production of 2,4,5-trichlorophenol and the subsequent production and
    use of 2,4,5-T acid and esters. The first commercial production of
    2,4,5-T in the United States was in 1944, and the use of 2,4,5-T
    herbicides increased in the 1940s and 1950s. However, the problem of
    dioxin contamination in 2,4,5-T was not recognized until 1957 (Kimmig
    & Schulz, 1957a, b).

         During the normal production of 2,4,5-T, the heaviest exposure to
    TCDD is during purification steps. The residues are far more
    contaminated than the purified products. Only limited information is
    available on the levels of TCDD contamination of products prepared
    prior to the 1970s, and absolutely no information is available on the
    dioxin levels in the corresponding residues. Consequently it is a
    difficult task to estimate the levels of occupational and general
    population exposures during the period prior to 1970.

        Table 25. Levels of PCDDs and PCDFs in samples of crab hepatopancreas
    from the west coast of Swedena
                                                                               

                                   Crab                Hepatopancreas
                                Idefjorden       Grebbestad       Vrfjord
                                  (pg/g)           (pg/g)           (pg/g)
                                                                               
        2,3,7,8-tetraCDF           31                47              590
          Total tetraCDFs          90               114              800

        2,3,7,8-tetraCDD           17                17              170
          Total tetraCDDs          17                17              170

      1,2,3,7,8-pentaCDFb           6                 7.6             45
      2,3,4,7,8-pentaCDF           44                50              130
          Total pentaCDFs         130               150              490

      1,2,3,7,8-pentaCDD           13                11               28
          Total pentaCDDs          86                76              270

    1,2,3,4,7,8-hexaCDFc           12                16               50
    1,2,3,6,7,8-hexaCDF             3                 5               10
    1,2,3,7,8,9-hexaCDF             3                 3               11
    2,3,4,6,7,8-hexaCDF            16                18               63
          Total hexaCDFs           70                88              280

    1,2,3,4,7,8-hexaCDD             8                 5               14
    1,2,3,6,7,8-hexaCDD            26                18               71
    1,2,3,7,8,9-hexaCDD             3                 4                7
          Total hexaCDDs          154               170              465

          Total heptaCDFs          23                28               90

          Total heptaCDDs          32                30               85

                octaCDF           < 1               < 1              < 2

                octaCDD           < 1               < 1              < 2
                                                                               

    a     From: Rappe et al. (1987).
    b     Not separated from 1,2,3,4,8-pentaCDF.
    c     Not separated from 1,2,3,4,7,9-hexaCDF.
    
        Table 26. TCDD content of the livers of farm animals from Seveso
    contaminated zones and surrounding areas (1976-1979)a
                                                                               
    Animal           Number         TCDD-containing         TCDD maximum
                   of samples           samples             level (ng/g)
                                                                               
    Rabbitsb         698               433                     633
    Poultry           83                35                      24
    Cattle            43                21                      94
    Horses            12                 2                      88
    Pigs              13                 0                       -
    Goats             25                17                       1
    Cats               1                 0                       -
                                                                               

    a  From: Pocchiari et al. (1983).
    b  Figures include rabbits kept in the special test plots on
         contaminated ground for experimental purposes.
    
        Table 27. CDD analyses of wildlife from Seveso contaminated zones and
    surrounding areas (1976-1979)a
                                                                                  
    Animal             Tested organs           Number of         Maximum level
                   and number of samples    TCD-containing          of TCDD
                                                samples             (ng/g)
                                                                                  
    Rabbits             6 (liver)                  4                  13
    Field mice         14 (whole body)            14                  49
    Rats                1 (pool-4 livers)                             28
    Earthworms          2 (pool)                                      12
    Frogs               1 (liver)                                      0.2
    Snakes              1 (liver)                                      3
                                                                               

    a     From: Pocchiari et al. (1983).
        
    Table 28. Analytical results for 2,3,7,8-tetraCDD residuesa
                                                                               
    Sections of 11      No. of deer  No. of   Concentration range  Limit of
    deer in study         samples   positive   of TCDD detected    detection
                         analyzed    samples       (ng/kg)b          range
                                                                   (ng/kg)b
                                                                               
    Muscle                 11          3          12 - 27           0.5 - 5
    Adipose tissue         10          8           3 - 12             1 - 3

    Table 28. (cont'd)  Analytical results for 2,3,7,8-tetraCDD residuesa
                                                                               
    Sections of 11      No. of deer  No. of   Concentration range  Limit of
    deer in study         samples   positive   of TCDD detected    detection
                         analyzed    samples       (ng/kg)b          range
                                                                   (ng/kg)b
                                                                               

    Liver                  11          4           2 -  5           0.4 - 4
    Bone marrow             5          0           ND                 1 - 3
                                                                               

    a    From: Harless et al. (1983). 
    b    Results corrected for efficiency of sample preparation.
         ND = not detected.
    
    4.4.4.1  Adipose tissue

         Gross et al. (1984) reported a study in which 30 coded samples of
    adipose tissue from Viet Nam veterans were analyzed for TCDD. The TCDD
    levels found for two of the three heavily exposed men were 99 pg/g and
    63 pg/g, which is higher than for the other Viet Nam veterans or for
    the controls (all well below 15 pg/g). Only one single isomer of
    tetraCDDs was found, and it was assumed that this was the
    2,3,7,8-isomer. The data in this study has also been discussed by
    Young et al. (1983). These authors concluded that the levels do not
    correlate well with known exposure data or with health status.

         Rappe et al. (1984) reported the presence of 2,3,7,8-substituted
    PCDDs and PCDFs in samples of human adipose tissue from Northern
    Sweden. A series of reports presented during the period 1984-1986
    confirms these observations and it has been clearly shown that there
    is a background of 2,3,7,8-substituted PCDDs and PCDFs in the general
    population in the industrialized part of the world. Most of these
    reports are lacking data on how the sampled people were selected and
    possible exposure to PCDDs and PCDFs. Consequently, these studies
    might not be representative. A series of earlier studies failed to
    detect these background levels due to insufficiently low detection
    levels. The Swedish study included 31 people, of which 18 were exposed
    to phenoxy esters and 13 were nonexposed. The group included 17 cancer
    patients and 14 non-cancer patients. The different groups were matched
    against each other. No difference in the levels, isomer patterns, or
    ranges could be found between these subgroups (Nygren et al., 1986).
    The mean values for these 31 people are given in Table 29.

         Schecter et al. (1986a) reported the mean levels of PCDDs and
    PCDFs in 46 samples of adipose tissue collected in Canada and in 8
    samples from the USA (see Table 29). The Canadian samples were taken
    from people who had died in 1976 from car accidents, drownings,

    trauma, and suicide. The samples included all ages and both sexes and
    came from all over the country. The USA samples, (1983-1984), were
    taken from biopsies from New York State residents during the course of
    normal medical procedures, and also from autopsies. Table 29 also
    includes the PCDD and PCDF levels in adipose tissue samples from Viet
    Nam (Schecter et al., 1986b) and from cancer patients in Japan (Ono et
    al., 1986).

         It is interesting to note the similarity between isomers present,
    levels of isomers, isomeric patterns, and congener profiles in samples
    collected from the general population in industrialized countries on
    three continents. The profile of the PCDD isomers shows increasing
    levels with an increasing number of chlorine atoms; the level of OCDD
    is 230-900 pg/g. On the other hand, the profile of PCDFs shows a
    maximum for 2,3,4,7,8-penta- or 1,2,3,6,7,8-hexaCDF. The difference in
    levels found between samples from South and North Viet Nam may be
    explained by spraying during the war in the 1960s and by the
    difference in industrial activities between the two parts of the
    country.

         Four samples of adipose tissue taken from German workers exposed
    to TCDD in the early 1950s have also been analyzed (Rappe et al.,
    1987). In spite of the fact that these workers were exposed more than
    30 years before collecting the samples, enhanced levels of
    2,3,7,8-tetraCDD could be identified, but the levels of the other
    PCDDs and PCDFs seem to be in the normal range (Table 29, last
    column).

         Patterson et al. (1986) studied the levels of 2,3,7,8-tetraCDD in
    the adipose tissue of 39 exposed people and 57 controls in Missouri,
    USA. The exposed group had subgroups of recreational, residential, and
    occupational exposure. All persons in both the exposed and control
    groups had detectable levels of 2,3,7,8-tetraCDD in their adipose
    tissue. Nineteen of the 39 exposed people had measurements higher than
    the highest level in the control group and six of the exposed people
    had levels greater than 100 ng/kg, which was five times higher than
    the highest control (Table 30).

         Ryan et al. (1986) analyzed autopsy tissue samples that were
    collected from three subjects who died in New York State, USA, from
    natural causes. The tissue types were: fat (both abdominal and
    subcutaneous), adrenal, bone marrow, liver, muscle, spleen, kidney,
    and lungs. As far as could be ascertained, no subjects had known
    abnormal exposure to PCDDs or PCDFs, yet these chemicals were found in
    all tissues analyzed. The highest concentrations of all PCDDs and
    PCDFs were found in adipose tissue. In individual tissues of the three
    subjects, the levels of individual congeners detected were within a
    narrow range, with much higher levels of the higher chlorinated PCDDs
    and PCDFs (e.g. hepta- and octa-CDD). In adipose tissue, the level of
    2,3,7,8-tetraCDD was 3.7-8.4 ng/kg and that of 2,3,4,7,8-pentaCDF

    5.2-13 ng/kg, while octaCDD ranged between 430 and 700 ng/kg. No major
    differences were seen between abdominal and subcutaneous fat samples
    or between these two types and perirenal fat when the lower lipid
    content of the latter was considered. Smaller concentrations of PCDDs
    and PCDFs were measured (on a wet weight basis) in decreasing order:
    adrenal, bone marrow, liver, muscle, spleen, kidney, and lungs.

         The high levels found in exposed Viet Nam veterans and German
    workers indicate a very slow excretion rate or metabolism of TCDD in
    humans. This indicates a dramatic difference between man and rodents;
    in the latter the half-life of TCDD is reported to be in the range of
    a few weeks.

    4.4.4.2  Blood plasma

         Analysis of blood plasma has been used to evaluate occupational
    exposure to PCDDs and PCDFs, which can occur during the production or
    use of 2,4,6-tri-, 2,3,4,6-tetra-, and pentachlorophenol. Rappe et al.
    (1983) investigated such exposure through the analysis of blood plasma
    of exposed workers and unexposed controls. Good correlations were
    found between the plasma levels and:

    (a)  the nature of exposure - dermal contact with liquids resulted in
         higher levels than inhalation of contaminated dust;

    (b)  the duration of exposure - higher levels for longer exposure
         times.

    The isomers present in the formulations used could also be found in
    the blood plasma.

         Kochman et al. (1986) studied a group of people exposed to PCBs
    and PCDFs after a transformer fire in Rheims, France in 1985. Low
    levels (1-25 pg/g) of 2,3,7,8-substituted penta-, hexa-, hepta-, and
    octaCDFs were found in the blood plasma of these people, and there was
    a slight variation in the T-lymphocytes cells.

         Kahn et al. (1986) measured the levels of PCDDs and PCDFs in the
    blood plasma from 10 heavily exposed Viet Nam veterans and their 17
    controls. The levels of TCDD was much higher in the exposed men than
    in their controls.



    
    Table 29. Levels of PCDDs and PCDFs in human adipose tissue (ng/kg wet weight)

                                                                                                                             
                                Sweden       USA/NYf       Canadaf       Japanf        N Viet Namff     FRG
            Isomer              n=31a        n=8b          n=46b         n=13c         n=9d          n=15d           n=4e


           2,3,7,8-tetraCDD       3            7.2           6.4 (25)      9 (12)       < 2            28 (12)       150
         1,2,3,7,8-pentaCDD      10           11.1          10 (46)       15 (13)       < 2            15 (14)        19.2
       1,2,3,6,7,8-hexaCDD       15           96            81 (46)       70 (12)        11 (6)       100 (15)        77
       1,2,3,7,8,9-hexaCDD        4           NA            NA            12 (10)        NA            NA              9.4
     1,2,3,4,6,7,8-heptaCDD      97          164           135 (46)       77 (12)        28 (6)       178 (15)        56
                   octaCDD      414          707           830 (46)      230 (12)       104 (8)      1256 (15)       267

           2,3,7,8-tetraCDF       3.9         NA            NA             9 (13)        NA            NA              0.9
         2,3,4,7,8-pentaCDF      54           14.3          15 (46)       25 (13)        13 (7)        21 (15)        44
       1,2,3,4,7,8-hexaCDF        6           NA            NA            15 (11)        NA            NA             10.0
       1,2,3,6,7,8-hexaCDF        5           31.3          16 (34)       14 (11)        13 (7)        58 (15)         6.7
       2,3,4,6,7,8-hexaCDF        2           NA            NA             8 (3)         NA            NA              3.8
     1,2,3,4,6,7,8-heptaCDF      11           16.5          30 (44)       NA              7 (3)        29 (15)        19.5
                   octaCDF        4           NA            NA            NA             NA            NA              1
                                                                                                                             

    n   = number of tissue samples.
    NA  = not analyzed.
    a   Nygren et al. (1986).
    b   Schecter et al. (1986a).
    c   Ono et al. (1986).
    d   Schecter et al. (1986b).
    e   Refers to occupationally exposed workers. Rappe et al. (1987).
    f   Mean values of positives. Number of positives within brackets.

        Levels below the detection level (< 1.0 ng/kg) not included.

    Table 30. Comparison of Levels of 2,3,7,8-Tetrachlorodibenzo-p-dioxin (ng/kg) in Adipose Tissue of Exposed and Control
    Groupsa

                                                                                                                             

                                                                       Exposed
                                                                                                                             
        Variable             Controls       Total          Recreational   Residential    Occupational
                                                                                                     
    Number of subjects       57             39              8             16              15
    Arithmetic mean           7.4           79.7           90.8           21.1           136.2
    Median                    6.4           17.0           23.5           14.5            24.7
    Range                     1.4-20.2       2.8-750        5.0-577        2.8-59.1        3.5-750
    Geometric mean            6.4           21.8           24.8           15.3            29.8
    Mean age, years (SD)     52.6 (15.7)    44.3 (13.7)    42.1 (14.7)    39.7 (14.9)     50.3 (9.8)
    % Men                    35.1           61.5           37.5           43.8            93.3
                                                                                                                             


    a     From: Patterson et al. (1986).

    


    5.  ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

    5.1  Air

         Owing to sampling and analytical problems, very few data are
    available on the levels of 2,3,7,8-TCDD and other PCDDs and PCDFs in
    normal urban air.

         Rappe & Kjeller (1987) reported the levels of PCDDs and PCDFs in
    total air samples collected in Hamburg, FRG, and samples of air
    particulates collected in Sweden (Table 31). Sample 1 was collected on
    the outskirts of Hamburg (representing an urban area), sample 2 in a
    traffic tunnel, sample 3 downwind from a MSW incinerator, and sample
    4 in the vicinity of a dumpsite and metal refinery. PCDDs and PCDFs
    were found in all samples. Lower levels of PCDDs and PCDFs were found
    in air particulates than in total air samples (Table 31). Sample 5 was
    taken when "clean air" was blowing into a street in Gothenburg,
    Sweden, and sample 6 was taken in the same street during an inversion
    situation. Samples 7 and 8 were taken at a rural research station
    outside Gothenburg when the air was blowing from the sea (sample 7) or
    from Gothenburg (sample 8). The isomeric pattern found in these
    samples were very similar (Rappe & Kjeller, 1987).

         Airborne dust was monitored in 1977 in the Seveso area to
    evaluate the possibility that 2,3,7,8-tetraCDD-contaminated particles
    might have drifted outside the contaminated areas. A high-volume
    sampling technique was used. When pooled particulate samples were
    analyzed, levels of 0.17-0.50 pg TCDD/m3 were reported (Wipf et al.,
    1982).

         The atmospheric concentrations of 2,3,7,8-tetraCDD near two
    hazardous waste sites have been monitored. In one study, US EPA (1982)
    failed to detect (detection limit: 1-20 pg/m3) any 2,3,7,8-TCDD in the
    atmosphere at the Love Canal (New York, USA) area. In another study of
    a waste disposal site (near Jacksonville, Arkansas, USA), Thibodeaux
    (1983) reported an average concentration of 1100 pg of 2,3,7,8-TCDD/g
    in two air particulate samples collected near the disposal site.

         Rappe et al. (1985c) analyzed indoor air samples for PCDFs
    resulting from fires and explosions in PCB-filled electrical
    equipment, and in an industrial situation (locomotive shop) (Table
    32).

         O'Keefe et al. (1985) analyzed air samples collected in an office
    building in Binghamton, New York, USA, after a transformer accident in
    the basement in February 1981. The samples were collected after a
    primary clean-up and the values are given in Table 33.

    
    Table 31. Levels of PCDDs and PCDFs in samples of total air and air
    particulatesa

                                                                                

      Total air samples                       Air particulates
                            1      2      3      4      5      6      7      8
                            pg/m3  pg/m3  pg/m3  pg/m3  fg/m3  fg/m3  fg/m3  fg/m3
                                                                                

    2,3,7,8-tetraCDFb       0.04   0.72   0.38   0.18    30     240     5     62
    Total tetraCDFs         0.36   6.2    4.9    3.3    320    2000    54    490

    2,3,7,8-tetraCDD        0.02   0.06   0.02   0.08     3       9   < 1      5
    Total tetraCDDs         0.10   0.22   0.21   1.5    150     350     9    130

    1,2,3,7,8-pentaCDFc     0.04   0.36   0.42   1.0     39     190     7     58
    2,3,4,7,8-pentaCDF      0.04   Inte   0.43   1.2     51     240     6     69
    Total pentaCDFs         0.51   4.1    5.0   10      470    2500    85    610

    1,2,3,7,8-pentaCDD    < 0.02   0.28   0.22   0.6     17      66     5     35
    Total pentaCDDs         0.07   1.3    2.4    5.0    200     840    31    280

    1,2,3,4,7,8-hexaCDFd    0.03   0.13   0.27   1.1     23     100     8     38
    1,2,3,6,7,8-hexaCDF     0.03   0.15   0.24   1.4     20      78     8     33
    1,2,3,7,8,9-hexaCDF   < 0.01 < 0.05 < 0.02   0.33     4      17     3     14
    2,3,4,6,7,8-hexaCDF   < 0.01 < 0.05   0.12   0.80    10      84     7     32
    Total hexaCDFs          0.18   1.1    2.2    9.5    180     800    70    310

    1,2,3,4,7,8-hexaCDD   < 0.08 < 0.17   0.19   1.0      3      19   < 1      7
    1,2,3,6,7,8-hexaCDD     0.23   0.66   0.71   2.2     11      46     4     14
    1,2,3,7,8,9-hexaCDD   < 0.08 < 0.17   0.36   5.2      6      92     5     32
    Total hexaCDDs          0.74   2.7    5.3   24.     100     520    32    190

    Total heptaCDFs         0.10   1.2    2.0    5.     200    1100   120    500

    Total heptaCDDs         0.60   3.4    5.3   15.     380    2900   140   1000

    OctaCDF               < 0.11 < 1.0    0.78   7.0    150     480   100    440

    OctaCDD                 0.37   6.4    7.4   40.0    290    1900    64    540
                                                                                

    a  From: Rappe & Kjeller (1987).
    b  Not separated from 2,3,4,8-tetraCDF.
    c  Not separated from 1,2,3,4,8-pentaCDF.
    d  Not separated from 1,2,3,4,7,9-hexaCDF.
    e  Int = Interferences.
    
    5.2  Water and Leachate

         Shadoff et al. (1977) failed to identify 2,3,7,8-tetraCDD in
    water from areas in the USA where 2,4,5-T herbicides had been used.

         In the 2,4,5-T plant in Jacksonville, Arkansas, USA, Thibodeaux
    (1983) could not detect any 2,3,7,8-tetraCDD in the creek water (no
    limit of detection levels given).

         Since August 1976, a number of tests have been periodically
    conducted in Seveso on streams running through the affected area, as
    far south as the River Lambro, with consistently negative results.
    During the same period sediment samples were taken from Torrents,
    Certesa, and Seveso. Positive results of the order of 1 pg/g were
    obtained within the first few kilometers downstream from their
    confluence, but further downstream results were negative. The
    intensive rainfalls after the accident caused the Seveso to repeatedly
    overflow its embankments at the point of entry into Milan, thus
    depositing silt on adjacent areas. Tests conducted to determine TCDD
    in these silts yielded negative findings for the first four floods
    while the fifth flood yielded positive findings (pg/g). Since August
    1976, the monthly determinations conducted on pipeline and ground
    waters have consistently yielded negative results, even when the
    analytical detection threshold was as low as 1 pg/litre (parts per
    quadrillion) (Pocchiari, 1983).

         During 1983 and 1984, the Dow Chemical Company conducted a study
    to determine the 2,3,7,8-tetraCDD contamination at its plant in
    Midland, Michigan, USA. It was estimated that 0.6 g of
    2,3,7,8-tetraCDD was being emitted per year in 2.5x107 m3 of
    wastewater effluent (Lamparski et al., 1986).

         The Ontario Ministry of Environment has included PCDDs and PCDFs
    in its Drinking Water Surveillance Program for the St Clair/Detroit
    River area (Ontario, 1986). No 2,3,7,8-tetraCDD has been found in any
    sample of raw or treated water. Unspecified congeners of PCDDs and
    PCDFs have been found mainly in raw water, octaCDD being the most
    frequent congener found. The highest value reported was 1.1 pg
    octaCDD/litre raw water in Amherstburg. The octaCDD level in the
    treated water was below the detection level of 0.01 pg/litre.

         Gtz (1986) reported on levels of PCDDs and PCDFs in the oily
    leachate from a sanitary landfill in Georgswerder, Hamburg, FRG (table
    34).

    5.3  Soil and Sediment

         The levels of 2,3,7,8-tetraCDD in point source after improper
    disposal of industrial waste are discussed in section 3.4.2.

        Table 32. Analyses of PCDFs in air samples (pg/m3)a
                                                                               
                          total
    Sample                tetra-  2,3,7,8-   penta-    hexa-   hepta-   octa-
                           CDF    tetraCDF    CDF       CDF     CDF      CDF
                                                                               
    Surahammar            < 20     < 2        < 10      < 10    < 10     < 10
    (during cleaning)

    Surahammar            < 10     < 2.5      < 10      < 10    < 10     < 10
    (after cleaning)

    Railway locomotive     500       50         50        30      20       20
    (during cleaning
    operations)
                                                                               

    a     From: Rappe et al. (1985c).
    
        Table 33. Concentrations of PCDFs in air samples collected on various
    floors of a Binghamton, New York (USA) office after primary cleanup

                                                                                     
                                       Analytical results (pg/m3)
                                                                                     

    Floor/sample typea     2,3,7,8-        Total           Penta-       Hexa-
                           tetraCDF      tetraCDFs          CDFs        CDFs
                                                                                     
      3                      16             151              43
      5                      11             126              30          2.0
      5 (NE)                 20             195              60          8.7
      7                      11             121              36
      9 volatiles            14             140              42
      9 particulates          1.8             4.8             4.7
      9 (SE) volatiles       13             146              31          3.7
      9 (SE) particulates     0.8             3.9             3.2
     11                      23              76              16
     11 (SE + NW)            16             133              19
     14                      11              92              21
     14 (NE)                 14             185              13
     16                      16             118              21
     17 volatiles            12              79              24
     17 particulates          0.8             3.9            NDb
     17 volatiles             9              59               6.6
     17 particulates          0.9           NDb               2.9
                                                                               

    Table 33. cont'd

    a     Abbreviations in parentheses designate sampling location on the floor,
          e.g., SE = south-east corner. Unless otherwise specified samples were
          collected in the north-west corner and analyzed as combined particulates
          and volatiles.

    b     ND = not detected.
    

        Table 34. Levels (ng/g) of the 2,3,7,8-substituted PCDDs and PCDFs in
              leachate from a sanitary landfilla
                                                                               
              Isomer      Concentration  Isomer                   Concentration
                                                                               
         2,3,7,8-tetraCDD      60        2,3,7,8-tetraCDF           9
       1,2,3,7,8-pentaCDD      28        1,2,3,7,8-pentaCDFb      322
     1,2,3,4,7,8-hexaCDD      476        2,3,4,7,8-pentaCDF       261
     1,2,3,6,7,8-hexaCDD     1440        1,2,3,4,7,8-hexaCDF      748
     1,2,3,7,8,9-hexaCDD      310        1,2,3,6,7,8-hexaCDF      336
                                         1,2,3,7,8,9-hexaCDF      558
                                         2,3,4,6,7,8-hexaCDF      114
                                                                               

    a     From: Gtz (1986).
    b     Overlapping isomer: 1,2,3,4,8-pentaCDF.
    
         Analytical results of the 1976-1977 survey of Zones B and R in
    Seveso were discussed by Pocchiari (1983). TCDD levels in Zones B and
    R were, in general, considerably lower than those in Zone A. In fact,
    most TCDD levels were lower than 50 g/m2 in Zone B and 5 g/m2 in
    Zone R. In 1980, a large part of Zone R was remonitored to evaluate
    the persistence of TCDD in the soil. This zone had been ploughed and
    worked since 1978. A comparison, as well as a statistical evaluation,
    of the relevant data indicated a significant decrease (40%) in the
    geometric mean level of TCDD in the soil of Zone R.

         In 1980 and 1981, soil samples from ten sites in Zone R and five
    sites outside Zone R were analyzed using a high resolution GS-MS
    system to establish whether other isomers of 2,3,7,8-tetraCDD were
    also present. A significant percentage decrease in tetraCDDs could be
    accounted for by two isomers (1,3,6,8-tetraCDD and 1,3,7,9-tetraCDD)
    present in the majority of the samples tested. These two isomers were
    not related to the chemical accident at the factory.

    
    Table 35. PCDD contamination in soil from Zone R in Seveso (1981)
    (values in ng/kg)a
                                                                               

    Sample    2,3,7,8b  TCDDc   Penta-    Hexa-     Hepta-    Octa-       Total
              tetraCDD           CDDs      CDDs      CDDs      CDD        PCDDs
                                                                               
      S1         0.8     0.3     0.4        6.0      1.4        1.7        10.6
      S2         3.4     1.0     0.7        9.5      2.1        2.2        18.9
      S3         4.0     1.9     1.2        8.2      8.6       27.0        50.9
      S4         2.3     0.8     0.6       10.2      2.1        2.0        18.0
      S5       < 0.1   < 0.3     0.5        9.5      1.9        1.4        13.7
      S6         6.3     1.5     1.1       10.4      2.6        1.3        24.8
      S7         1.7     1.0     0.8       12.4      1.9        1.8        19.6
      S8         2.2     0.4     0.8        8.8      1.8        0.8        14.8
      S9         1.0     2.8     2.3       21.2      9.6       13.5        50.4
                                                                               

    a     From:  Wipf & Schmid (1983).
    b     Probably related to the accident.
    c     Total levels of isomers other than 2,3,7,8-tetraCDD,
          probably not related to accident.
    
         Wipf & Schmid (1983) reported the presence of PCDDs other than
    2,3,7,8-tetraCDD in the soil from Zone R (Table 35). They suggest that
    a municipal incinerator and the burning of wood shavings treated with
    chlorinated phenols could be the source of the other PCDDs.

         Nestrick et al. (1986) reported the levels of 2,3,7,8-tetraCDD in
    soil samples collected from industrialized areas of US cities. They
    observed a widespread occurrence of 2,3,7,8-tetraCDD in urban soils,
    with levels of 1-10 ng/kg, and suggested that local combustion
    sources, including MSW and industrial incinerators, were the probable
    origin.

         McLaughlin & Pearson (1984) measured soil concentrations of PCDDs
    and PCDFs in the vicinity of a municipal refuse incinerator in
    Ontario, Canada. Urban and rural control locations were also sampled.
    All soil samples (14) had detectable quantities of at least one of the
    five PCDD congener classes (tetraCDD to octaCCD) tested for, whereas
    eight samples contained detectable levels of one or more of the five
    PCDF congener groups (tetraCDF to octaCDF). The levels ranged from
    non-detectable (0.003 - 0.008 ng/g) to 3.5 ng/g (octaCDD); only one
    site had a measurable quantity (0.007 ng/g) of tetraCDD in the soil.
    The most abundant PCDD or PCDF congener was octaCDD, which had similar
    levels whether samples were taken close to or remote from the
    incinerator. Similarly, no concentration gradients, relative to
    distance from the incinerator, were apparent for any of the other
    PCDDs or PCDFs.

         Soil samples near a chemical waste incinerator in Scotland have
    also been analyzed for PCDDs and PCDFs, together with samples from
    control locations (Edulgee et al., 1986). Detectable levels of all
    PCDFs or PCDFs that were examined were found in each of the soil
    samples (13). Levels found ranged from 1.2 ng/kg (2,3,7,8-tetraCDD) to
    1900 ng/kg (total hexaCDF). No consistent pattern was observed to
    differentiate levels found in control samples from levels in soil near
    the incinerator.

         These studies from widely separate areas of the world support the
    suggestion that diffuse combustion sources are the major source of
    PCDDs and PCDFs in the soil.

         Rappe & Kjeller (1987) analyzed soil samples from various parts
    of Europe (Table 36). They represent rural areas (samples 1, 2, and 3)
    as well as more industrialized areas (samples 4 and 5). PCDDs and
    PCDFs could be identified in all samples. The 2,3,7,8-tetraCDD
    concentration was below the detection level in the soil samples from
    the non-industrialized areas. Trapped sediments from the archipelago
    of Stockholm, Sweden, were also analyzed. The samples were collected
    in the inner (sample 6), middle (sample 7), and outer archipelago
    (sample 8). Levels decreased with increasing distance from the city of
    Stockholm. The isomeric patterns for tetra- and pentaCDF isomers are
    very similar to those found for samples of total air and air
    particulates (section 6.1). A sediment sample from the mouth of River
    Viskan, Sweden, was also analyzed (sample 9). A slight difference in
    congener profile was found between this sample and the sediments from
    the archipelago of Stockholm.

         Czuczwa & Hites (1985) found PCDDs and PCDFs in sediment samples
    from several locations in Saginaw River and Bay, and southern Lake
    Huron, levels ranging from 100 ng/g in urban areas to 100 ng/kg at
    remote sites. Although no isomers were identified, the analytical
    profiles in the sediments followed closely those found in combustion
    samples, suggesting that combustion is the major source of PCDDs and
    PCDFs found in the sediments. Analyses of sediment cores showed a
    dramatic increase in the PCDD and PCDF concentrations at a depth
    corresponding to approximately the year 1940, and levels remained high
    up to the present. There is no good correlation between the trend in
    these levels and the trend for coal burning in the United States.
    However, the levels in the sediments correlate with the production and
    use of chlorinated aromatic compounds within this area of the Great
    Lakes.

    5.4  Food

    5.4.1  Meat and bovine milk

         The levels of PCDDs and PCDFs in fish and other seafood are
    discussed in section 4.4.2.



        Table 36. Levels (pg/g) of PCDDs and PCDFs in samples of sediments and soila
                                                                                                                           
                                          Soil samples                                     Sediments
                             1        2        3        4        5              6         7         8        9
                                                                                                                           
    2,3,7,8-tetraCDFb        2.9      1.6      1.1      34        38             30        17       14        1.6
    Total tetraCDFs          9.3      7.7     11       320       370            290       150      120       24

    2,3,7,8-tetraCDD       < 2.0    < 2.1    < 0.2       2.4       0.8            2.4       2.0    < 2.0       0.2
    Total tetraCDDs           -        -       3.2      55.5      11.2           69        21       23        6.4

    1,2,3,7,8-pentaCDFc      2.5      1.6      0.5      17        31             16         8.6      8.5      1.3
    2,3,4,7,8-pentaCDF       0.8      1.0      0.6      23        65             20        14       16        1.7
    Total pentaCDFs         14       13        6.7     200       450            260       140      130       30

    1,2,3,7,8-pentaCDD     < 2.0    < 2.0    < 0.1      18        34              7.6       5.2      5.5      0.9
    Total pentaCDDs           -        -       4.6     220       270            230        99       86       13

    1,2,3,4,7,8-hexaCDFd     3.8      2.2      0.9      30        45             16        10        8.8      1.9
    1,2,3,6,7,8-hexaCDF      1.8      1.5      0.4      11        25             12         7.1      5.9      1.2
    1,2,3,7,8,9-hexaCDF      1.0      0.9      4.3     110      1100              5       < 1      < 1        2.0
    2,3,4,6,7,8-hexaCDF      1.9      1.0      0.7      26        57             16        31       23        1.6
    Total hexaCDFs          16       12       11       270      1900            250       220       92       44

    1,2,3,4,7,8-hexaCDD    < 2      < 2      < 0.1      13        28              1.6       0.8      1.0      1.6
    1,2,3,6,7,8-hexaCDD    < 2      < 2      < 0.1      19        64             48         2.0      2.0     10
    1,2,3,7,8,9-hexaCDD    < 2      < 2      < 0.1       6.2      19              2.5       0.9      1.0      4.3
    Total hexaCDDs            -        -       4.7     200       330             49        16       19       64
    Total heptaCDFs         22       14       18       260      4500           1300      1500      190      300
    Total heptaCDDs       < 10     < 10       17       370      1600           5700      1200      880      190
    OctaCDF                   -        -       5.7      68        71             39      < 20     < 20      330
    OctaCDD                   -        -      14       140       180           3100       510      260      900
                                                                                                                           
    a     From: Rappe & Kjeller (1987).
    b     Not separated from 2,3,4,8-tetraCDF, except in the case of sample 9.
    c     Not separated from 1,2,3,4,8-pentaCDF.
    d     Not separated from 1,2,3,4,7,9-hexaCDF.
    


         The US Environmental Protection Agency (US EPA) initiated a
    2,3,7,8-tetraCDD monitoring programme of beef fat samples taken from
    cattle that had grazed on rangelands known to have been treated with
    2,4,5-T. The analytical collaborators in this programme were Dow
    Chemical Co. (USA), Wright State University, and Harvard University.
    All the laboratories used mass spectroscopic techniques for
    quantification. Two different extraction techniques were used. All
    three laboratories analyzed control samples taken from cattle that had
    grazed on non-treated areas. Some of these control samples were spiked
    with known amounts of TCDD. All the controls were prepared by the US
    EPA (Firestone, 1978). Good agreement was found between the amounts of
    TCDD spiked into the control beef fat samples and the reported levels
    found, even down to TCDD levels of 10 ng/kg. The average reported TCDD
    level was 10 ng/kg; the amount actually added by the EPA was 9 ng/kg.
    Of a total of 34 analyses of controls to which no TCDD was added, in
    only one case was there a false positive report of TCDD (O'Keefe et
    al., 1977). Of 52 samples of beef fat from 2,4,5-T-treated rangeland,
    19 (37%) were reported by one or more laboratories to have TCDD. The
    average range of levels reported was 5-66 ng/kg, and the overall
    average was 7 ng/kg. If one considers only the 40 beef fat samples
    from areas receiving at least 1.1 kg of 2,4,5-T/ha, all 19 positive
    samples (48%) belong to this group, and the average reported TCDD
    level would be 9 ng/kg. The results indicated a consistent trend,
    relating the average reported TCDD level in beef fat to the intensity
    of the 2,4,5-T application to rangeland (O'Keffe et al., 1978;
    McKinney, 1978).

         None of the three collaborating laboratories used the most
    selective and sensitive analytical method now known (capillary glass
    column gas chromatography - high resolution mass fragmentography).

         Kocher et al. (1978) also analyzed specimens of fat taken from
    steers that had grazed on rangeland previously treated with 2,4,5-T
    herbicides. The limit of detection of TCDD (2.5 times peak to peak
    noise) was found to be in the 30-60 pg range (3-6 pg/g in beef fat
    using 10 gram samples). None of the sixteen samples analyzed in two of
    three studies revealed TCDD. In the third study, the animals were
    confined to a fenced pasture sprayed in its entirety with 2,4,5-T
    herbicides. The samples from three of the seven animals gave a
    positive response at the extremely low level of 3 to 4 ng TCDD/kg,
    which is at the detection limit; the highest reported value (without
    interfering components) was 13 pg/g. The level of 2,3,7,8-tetraCDD in
    the 2,4,5-T used was, however, unknown. Beck et al. (1986) analyzed
    seven randomly collected samples of cow's milk from different areas of
    the Federal Republic of Germany and one sample from the German
    Democratic Republic. The cow's milk was taken from road transport
    tankers. In all samples, 2,3,7,8-substituted PCDDs and PCDFs were
    found at low levels of pg/g on a fat weight basis. Detection limits

    were in the range 0.1-0.3 ng/kg. The levels of PCDDs and PCDFs found
    in the milk samples were lower than those measured by Rappe et al.
    (1987b) in cow's milk from Switzerland (Table 38). In addition, there
    was no evidence of high levels of the higher chlorinated congeners
    (e.g. hepta and octa- ) (see Table 37). Levels were much lower than
    those in human milk samples (see section 5.4.2).

         Rappe et al. (1987b) analyzed PCDDs and PCDFs in six samples of
    bovine milk from various locations in Switzerland. In all samples,
    2,3,7,8-substituted PCDDs and PCDFs were found at levels of pg/kg in
    whole milk (Table 38). However, the levels were lower in commercial
    milk samples than in samples collected directly from cows grazing in
    the vicinity of incinerators.

         When Ryan et al. (1985) analyzed PCDDs and PCDFs in chicken and
    pork samples in Canada, the incidence of positives for hexa-, hepta-,
    and octaCDD in selected samples of chicken fat was 50, 62, and 46%,
    with averages of 27, 52, and 90 ng/kg, respectively. Similar levels of
    hexa- and heptaCDFs were also found in some of these samples, but
    tetra- and pentaCDDs and tetra-, penta-, and octaCDFs were not
    detected. A comparison between the tissue analyses and those of the
    wood (treated with pentachlorophenol) used to house the animals showed
    a marked similarity (see Table 38), indicating that pentachlorophenol
    was the probable source of contamination of the food samples.

         Firestone et al. (1986) reported the analyses of various food
    items collected in the period beginning in 1979. Low levels (< 300
    ng/kg) of 1,2,3,4,6,7,8- and 1,2,3,4,6,7,9-heptaCDD were found in some
    samples of chicken, bacon, pork chops, and beef liver. HexaCDD was not
    found in any of the foods. Several beef livers had high levels of OCDD
    residues, the highest reported value being 3830 ng/kg. No PCDDs (at a
    detection limit of 10-40 ng/kg) were found in ground beef.

    5.4.2  Human milk

         Rappe et al. (1984) reported low levels of 2,3,7,8-substituted
    PCDDs and PCDFs in five samples of human milk from the Federal
    Republic of Germany (FRG) (Table 39).

         Table 39 also indicates the results of analyses of four samples
    of human milk from the Umea region of northern Sweden (Rappe, 1985),
    92 samples from Rheinland-Westfalia in the FRG (Furst et al., 1987),
    30 other samples from the FRG (Beck et al., 1987), and five samples
    from the Netherlands or Yugoslavia (Rappe et al., 1987).

         Van der Berg et al. (1986) also reported the levels of PCDDs and
    PCDFs in human milk samples from the Netherlands, but these levels
    were reported on milk basis, not on fat basis, and are therefore not
    included in Table 39.

         A comparison between the isomers, levels, isomeric pattern, and
    congener profiles found in human milk (Table 39) and in adipose tissue
    (Tables 29 and 30) shows a remarkable degree of similarity.

    5.4.3  Rice

         Rice from fields in Arkansas, Louisiana, and Texas, USA, treated
    at a maximum rate to give 2.52 kg 2,4,5-T/ha, were analyzed for
    possible 2,3,7,8-tetraCDD residues. A specification of 1 g
    2,3,7,8-tetraCDD/g 2,4,5-T was given in the report, but no analytical
    data were given for the herbicide. No 2,3,7,8-TCDD was detected in the
    rice (detection limit = 2-7 g/kg) and no 2,3,7,8-TCDD residues
    (detection limit = 2-10 g/kg) were found in 30 samples of rice
    purchased in retail stores throughout the USA (Jensen et al., 1983).

    5.5  Yusho and Yu-cheng Episodes

         In 1968, more than 1500 people in south-west Japan were
    intoxicated through consuming commercial rice oil accidentally
    contaminated by PCBs, PCDFs, and polychlorinated quarterphenyls
    (Masuda & Yoshimura, 1982; Masuda et al., 1985). In 1979, a similar
    episode occurred in central Taiwan, the number of persons involved
    here approaching 2000 (Chen et al., 1980; Chen et al., 1985). In the
    past, both accidents were referred to as Yusho episodes, but now the
    Taiwan episode has been renamed Yu-cheng.

         The Japanese rice oil contained more than 40 PCDF isomers (tri-
    to hexaCDFs) (Buser et al., 1978d), whereas the number of isomers in
    the Taiwanese oil seems to have been less (Chen & Hites, 1983). The
    toxic 2,3,7,8-substituted PCDFs were middle or minor constituents,
    about 10-15% of the total amount of PCDFs (Buser et al., 1978d; Masuda
    et al., 1985; Chen & Hites, 1983). The mean total consumption of PCDFs
    of the Yusho and Yu-cheng patients has been estimated to be 3.3-3.8
    mg/person (Hayabuchi et al., 1979; Chen et al., 1985), or a daily
    intake of total PCDFs of 0.9 g/kg body weight (Hayabuchi et al.,
    1979). The average intake of 2,3,7,8-substituted PCDFs was 90-135
    ng/kg body weight per day. The smallest amount of total PCDFs causing
    chloracne has been estimated to be 0.16 g/kg body weight per day
    (Hayabuchi et al., 1979) or 20-30 ng/kg per day of the
    2,3,7,8-substituted congeners.

         Analysis of liver samples taken from the Yusho patients about 18
    months after the exposure showed a dramatic decrease in the number of
    PCDF isomers. Apparently most of the PCDF isomers were metabolized or
    excreted during the period between exposure and sampling (Rappe et
    al., 1979). A comparison between the PCDF isomers found in the Yusho
    oil and the liver samples revealed an interesting relationship. Most
    of the isomers retained had lateral positions (2, 3, 7, and 8)
    substituted with chlorine; these isomers have the highest toxicity
    (Rappe et al., 1979).



    
    Table 37. PCDD and PCDF levels in samples of cow's milk (ng/kg on a fat weight basis)a

    SAMPLE
                                                                                                                           
                                     1         2         3         4         5         6         7         8
                                                                                                                           
          2,3,7,8-tetraCDF        < 0.1       0.29      0.28      1.1       1.0       1.4       1.4       0.27
          2,3,7,8-tetraCDD          0.33    < 0.2     < 0.2     < 0.2       ND      < 0.2       ND        ND
        1,2,3,7,8-pentaCDF          ND        0.4       ND        0.26      0.24      ND        0.39    < 0.2
        2,3,4,7,8-pentaCDF          1.3       0.91      1.1       1.6       1.5       1.3       2.9       0.8
        1,2,3,7,8-pentaCDD          1.0       0.72      ND        0.81      0.6       0.78      1.2     < 0.5
      1,2,3,4,7,8-hexaCDF           0.93      0.67      0.70      0.85    < 0.3       0.84      1.9       0.57
      1,2,3,6,7,8-hexaCDF           0.73      0.58      0.57      0.85    < 0.3       0.73      2.1       0.41
      2,3,4,6,7,8-hexaCDF           0.65      0.48      0.53      0.68      ND        0.64      1.8       0.37
      1,2,3,4,7,8-hexaCDD           0.33      0.34    < 0.3     < 0.3     < 0.3       0.36      0.33    < 0.3
      1,2,3,6,7,8-hexaCDD           1.3       1.2       1.0       0.82      0.32      1.7       1.9       0.80
      1,2,3,7,8,9-hexaCDD         < 0.3       0.34      0.36      0.39      ND        0.55      0.48    < 0.3
    1,2,3,4,6,7,8-heptaCDFb          < 0.5     < 0.5     < 0.5     < 0.5     < 0.5     < 0.5     < 0.5     < 0.5
    1,2,3,4,6,7,8-heptaCDDb          < 2       < 2       < 2       < 2       < 2       < 2       < 2       < 2
                  octaCDDb          <10       <10       <10       <10       <10       <10       <10       <10
                  octaCDFb          < 1       < 1       < 1       < 1       < 1       < 1       < 1       < 1


    a    From: Beck et al. (1987).
    b    Not significantly higher than blanks.
    ND   = not detectable.

    Table 38. PCDF and PCDD content of bovine milk from Switzerland. Results in ng/kg (ppt), whole milk basisa
                                                                                                                           
    Compound                                   Commercial                                      Incinerators
                                     1              2              3                  4              5              6
                                                                                                                           
           2,3,7,8-tetraCDF      < 0.028         < 0.035        < 0.021            < 0.022        < 0.032        < 0.028
         1,2,3,7,8-pentaCDF      < 0.020         < 0.022        < 0.021            < 0.020        < 0.036        < 0.032
         2,3,4,7,8-pentaCDF        0.084           0.066          0.069              0.43           0.22           0.23
       1,2,3,4,7,8-hexaCDF       < 0.020         < 0.026        < 0.017              0.13           0.06           0.084
       1,2,3,6,7,8-hexaCDF         0.028         < 0.018        < 0.021              0.19           0.095          0.059
       2,3,4,6,7,8-hexaCDF       < 0.020         < 0.018          ND                 0.28           0.12           0.049
                                                               (< 0.02)
     1,2,3,4,6,7,8-heptaCDF      < 0.12            ND             ND                 0.49           0.28         < 0.18
                                                (< 0.13)       (< 0.08)
                   octaCDF       < 0.20            ND             ND                 ND             ND           < 0.52
                                                (< 0.13)       (< 0.09)           (< 0.16)       (< 0.21)

           2,3,7,8-tetraCDD        ND              ND             ND                 0.049          0.038          0.021
                                (< 0.012)       (< 0.013)      (< 0.013)
         1,2,3,7,8-pentaCDD        ND              ND             ND                 0.25         < 0.086          ND
                                (< 0.04)        (< 0.08)       (< 0.06)                                         (< 0.1)
       1,2,3,4,7,8-hexaCDD      < 0.068            ND             ND                 0.23           0.14         < 0.14
                                                (< 0.1)        (< 0.06)
       1,2,3,6,7,8-hexaCDD      < 0.068            ND             ND                 0.29           0.16         < 0.21
                                                (< 0.1)        (< 0.06)
       1,2,3,7,8,9-hexaCDD      < 0.068            ND             ND                 0.17         < 0.080        < 0.11
                                                (< 0.1)        (< 0.06)
     1,2,3,4,6,7,8-heptaCDD     < 0.064          < 0.066        < 0.064              0.26         < 0.095          0.42
                   octaCDD      < 0.16           < 0.26         < 0.12               0.28         < 0.16           0.59
                                                                                                                           


    a    From: Rappe et al. (1987b).
    ND   = not detectable.

    
    Table 39. Levels of PCDDs and PCDFs found in human milk (ng/kg of fat weight).
                                                                                                                             
                                       Sweden    FRG       FRG       FRG       Netherlands    Yugoslavia
                                       n=4a      n=5b      n=92c     n=30d     n=3e           n=2e
                                                                                                                             
               2,3,7,8-tetraCDD          0.6       1.9     < 5        3.4       9.7          < 1.0
             1,2,3,7,8-pentaCDD          6.5      12.9      10.7      15        44              5.5
            1,2,3,4,7,8-hexaCDD          2.5       4.6       8.1      12        25              3.5
            1,2,3,6,7,8-hexaCDD         19        17.3      32.7      59       251             15
            1,2,3,7,8,9-hexaCDD          6.3       1.6       6.4      11        23             ND
         1,2,3,4,6,7,8-heptaCDD         59.5      72.8      49.9      61       130            106
                        octaCDD        302       434       181       530       744            106

               2,3,7,8-tetraCDF          4.2       5.4       2.6       2.5       2.8          < 1.0
             1,2,3,7,8-pentaCDF        < 1       < 1         1.8     < 1        ND             ND
             2,3,4,7,8-pentaCDF         21.3      36.4      22.9      20        79             25
            1,2,3,4,7,8-hexaCDF          4.7      11.4       8.2       8.5       8.9            3.7
            1,2,3,6,7,8-hexaCDF          3.4      10.2       6.6       7.8      10.3            3.6
            2,3,4,6,7,8-hexaCDF          1.4       4.3       3.3       3.0       6.4            1.3
         1,2,3,4,6,7,8-heptaCDF          7.4       9.2       6.4       8.5      39             ND
                       octaCDF           3.2       2.4      22.8     < 3        ND             ND

                                                                                                                             

    ND   = not detected; NA = not analyzed; n = number of samples.
    a    Rappe (1985).
    b    Rappe et al. (1984).
    c    Furst et al. (1987).
    d    Beck et al. (1987).
    e    Rappe et al. (1987) (pooled samples)
    


         Kunita et al. (1984) studied the blood levels of PCDFs in people
    involved in the Japanese and Taiwanese episodes. They used a
    non-isomer-specific analytical method, and reported higher blood
    levels in persons with severe dermal symptoms than in persons with
    light symptoms. The levels 2 years after exposure were lower than 0.5
    year after exposure; for the six persons studied the levels of total
    PCDFs decreased on average by 57%  12% (Kunita et al., 1984).

         The rate of excretion of these toxic PCDF isomers is very low.
    Rappe & Nygren (1984) could detect 2,3,4,7,8-pentaCDF in blood plasma
    from Yusho patients when the samples were collected 11 years after
    exposure. However, higher levels were found in blood from Yu-cheng
    patients one year after exposure; these analyses also showed a 15-20%
    reduction in levels in one year (Rappe, 1984).

    6.  KINETICS AND METABOLISM OF 2,3,7,8-TETRACHLORO-
    DIBENZO-P-DIOXIN (TCDD) AND OTHER PCDDs

    6.1  Uptake, Distribution, and Excretion

         Most of the available toxicokinetic data arise from studies of
    gastrointestinal exposure and oral or intraperitoneal administration.
    There has been one dermal study on rats, but no studies on exposure
    via the respiratory tract. Data on the gastrointestinal absorption,
    distribution, and elimination of TCDD in various species are
    summarized in Tables 40 and 41. Principle organ depots in all species
    studied are the liver and adipose tissue. In addition skin and muscle
    have been found to be organ depots in monkeys, and skin in
    guinea-pigs. In all species studied clearance of TCDD from the body
    follows apparent first-order kinetics.

    6.1.1  Studies on rats

         Male Sprague Dawley rats were dosed by gavage with 14C TCDD in
    acetone: corn oil (1:9), at a concentration of 50 g/kg body weight
    (Piper et al., 1973). Rats lost weight and their physical condition
    was generally poor, although no deaths occurred. Approximately 30% of
    the administered dose of radioactivity was eliminated in the faeces
    during the first 48 h, this being most probably unabsorbed TCDD. The
    total faecal TCDD content was equivalent to 53.2% of the dose over a
    21-day period. During this time 13.2% was eliminated in the urine and
    3.2% in the air. From these results the half-time for the clearance of
    TCDD was calculated as 17.4 ( 5.6) days. In another study, groups of
    two to three rats were killed at different time intervals after
    administration of the same dose of 14C-TCDD. The liver contained
    3.2, 4.5, and 1.3% of the administered dose per gram of tissue 3, 7,
    and 21 days, respectively, after dosing. The concentrations in adipose
    tissue at the same time intervals were 2.6, 3.2, and 0.4% of the dose
    per gram of tissue. Other tissues showed lower concentrations.

         In a study by Allen et al. (1975), male Sprague Dawley rats were
    given 14C-TCDD in a single dose of 50 g/kg body weight by stomach
    tube. Groups of five rats were sacrificed 1, 3, 5, 7, 14, and 21 days
    after dosing. The dose given resulted in marked liver hypertrophy,
    thymic regression, weight loss, and death in 50% of the animals within
    25 days. Twenty-five percent of the dose was eliminated within the
    first 3 days through the faeces. During the next 18 days, 1 to 2% of
    the dose was found in the faeces daily. The total amount in the faeces
    during the 21 days following administration was about 52%.



    
    Table 40.  Gastrointestinal absorption of TCDD
                                                                                                                             
    Species/Strain        Vehicle                     Dose                           % Absorption      Reference
                                                      (g/kg body weight)            (mean  SD)
                                                                                                                             
    Rats
        Sprague Dawley     acetone:corn oil (1:9)      50a                            70            Piper et al. (1973)
        Sprague Dawley     corn oil                    50a                          > 75            Allen et al. (1975)
        Sprague Dawley     diet                        7 or 20; for 42 days           50-60         Fries & Marrow (1975)
        Sprague Dawley     acetone:corn oil (1:24)     1a                             8411         Rose et al. (1976)
        Sprague Dawley     acetone:corn oil (1:24)     0.1 or 1.0; 5 days/week        8612         Rose et al. (1976)
                                                       for 7 weeks

    Mice
        ICR/Ha Swiss       ethanol:Tween 80:saline     135a                           27-28         Koshakji et al. (1984)
                           (1:10:89)

    Hamsters
        Golden Syrian      olive oil                   650a                         73.522.8       Olson et al. (1980a)
                                                                                                                             


    a     Single dose.

    Table 41.  Elimination of TCDD in different species
                                                                                                                             

    Species/Strain    Route/Vehicle        Dosea      Duration  Half-life       Eliminated radioactivity      Reference
                                           (g/kg     of        for elimination (% of administered dose)
                                           body       study     (days)
                                           weight)    (days)                    Faeces         Urine
                                                                                                                             
    Rats

    Sprague Dawley    oral/acetone:        50         21        17.4 5.6       53.2           13.2           Piper et al. (1973)
                        corn oil (1:9)
    Sprague Dawley    oral/corn oil        50         21        21.3 2.9       53.3           4.5            Allen et al. (1975)
    Sprague Dawley    oral/acetone:         1         22        31  6          NR             ND             Rose et al. (1976)
                        corn oil (1:24)
    Sprague Dawley    oral/acetone:         1b        49        23.7            NR             3.1 0.2 (m)   Rose et al. (1976)
                        corn oil (1:24)                                                        12.5 5.1 (f)  
    Sprague Dawley    ip/corn oil          400         7        NR              4.96 0.3      0.51 0.05     van Miller et al.
                                                                                                                (1976)
    Mice
    C57BL/6           ip/olive oil         10         30        11              59.3           20.5           Gasiewicz et al.
                                                                                                                (1983b)
    DBA/2             ip/olive oil         10         30        24.4            39.8           16.8           Gasiewicz et al.
                                                                                                                (1983b)
    B6D2F1            ip/olive oil         10         30        12.6            56.3           20.5           Gasiewicz et al.
                                                                                                                (1983b)
    ICR/Ha Swiss      oral/ethanol:        135        11        20              78             4              Koshakji et al.
                        tween 80:                                                                               (1984)
                        saline (1:10:89)
    C57BL/6 Ahb/Ahd ip/emulphor:           0.5        42        9.6             61.8           19.9           Birnbaum (1986)
                        ethanol:H2O
                        (1:1:18)
    C57BL/6 Ahd/Ahd ip/emulphor:           0.5        42        9.6             55.9           26.8           Birnbaum (1986)
                        ethanol:H2O
                        (1:1:18)
    DBA/2 Ahb/Ahd ip/emulphor:             0.5        42        10.8            71.5           13.6           Birnbaum (1986)
                        ethanol:H2O
                        (1:1:18)
                                                                                                                             

    Table 41 (contd).
                                                                                                                             
    Species/Strain    Route/Vehicle        Dosea      Duration  Half-life       Eliminated radioactivity      Reference
                                           (g/kg     of        for elimination (% of administered dose)
                                           body       study     (days)
                                           weight)    (days)                    Faeces         Urine
                                                                                                                             
    Mice (contd)

    DBA/2 Ahd/Ahd ip/emulphor:             0.5        42        10.8            76.1           11.1           Birnbaum (1986)
                        ethanol:H2O 
                        (1:1:18)

    Guinea-pigs

    Hartley           ip/olive oil         2          23                        33              2             Gasiewicz &
                                                                                                                Neal (1979)
    NR                oral/NR              NR         22        22-43           ND             ND             Nolan et al. (1979)
    Hartley           ip/olive oil         0.56       45        93.7  15.5     26.2 3.6      3.1 1.2       Olson (1986)

    Hamsters

    Golden Syrian     oral/olive oil       650        35        15.0 2.5       NR             NR             Olson et al.
                                                                                                                (1980a)
    Golden Syrian     ip/olive oil         650        35        12.0 2.0       50.02.7       34.65.4       Olson et al.
                                                                                                                (1980a)
    Monkeys

    Rhesus (adult)    ip/corn oil          400         7        NR              3.75           1.06           van Miller et al.
                                                                                                                (1976)
    Rhesus (infant)   ip/corn oil          400         7        NR              1.26           2              van Miller et al.
                                                                                                                (1976)
                                                                                                                             

    a     Single dose, unless otherwise stated.
    b     5 days/week for 7 weeks.
    ND   = not detectable, NR = not reported, m = males, f = females, ip = intraperitoneal.
    


         The authors concluded that the faecal content during the first 3
    days represented mainly unabsorbed TCDD and that apparently more than
    75% of the administered dose had been absorbed from the
    gastrointestinal tract. The radioactivity excreted daily through the
    urine ranged between 0.1 and 0.2% of the administered dose during the
    initial 12 days. Thereafter, the daily urinary radioactivity excretion
    increased from 0.25% of the administered dose to 0.43% by day 21. The
    total amount of 14C excreted through the urine over a 21-day period
    was about 4.5% of the dose. On the basis of the daily faecal and
    urinary excretion, the authors calculated a half-time of 21.3 ( 2.9)
    days. In the rats killed on days 1, 3, 5, 7, 14, and 21 after
    administration, the total liver content was 56 ( 5), 54 ( 14), 54 (
    6), 54 ( 8), 45 ( 5), and 24 ( 4)% of the administered dose,
    respectively. At all intervals, the levels found in the liver exceeded
    those found in other organs. On days 5, 7, and 14 about 90% of the
    total radioactivity in the liver was present in the microsomal
    fraction.

         Fries & Marrow (1975) fed Sprague Dawley male and female rats a
    diet containing 7 or 20 g 14C-TCDD/kg diet for 42 days. Thereafter
    all rats received the control diet for another 30 days. Two animals of
    each sex and TCDD dietary level were sacrificed at 14-day intervals.
    This treatment resulted in decreased food consumption, decreased
    weight gain, and increased relative liver weight among both males and
    females. The concentration of TCDD-derived radioactivity in the liver,
    which was the principal tissue depot of both males and females, was
    directly proportional to the dietary intake of 14C-TCDD. At the end
    of the 42-day feeding period, the 14C levels in male livers
    indicated TCDD contents of 5.8 and 15.9 g/kg of tissue for the lower
    and higher feed concentrations, respectively. The concentrations in
    the liver of female rats were similar. Analysis of the liver of rats
    killed 14 and 30 days after discontinuing TCDD exposure indicated a
    gradual decrease in TCDD liver concentration in both sexes. The
    steady-state body burden for TCDD was estimated to be 10-11 times the
    daily intake in both sexes. The whole-body and liver half-lives were
    calculated to be 12 and 11 days in males and 15 and 13 days,
    respectively, in females.

         Rose et al. (1976) estimated the absorption of a single non-toxic
    dose of 1 g 14C-TCDD/kg body weight in male and female Sprague
    Dawley rats to be 84  11% of the administered dose. The elimination
    of TCDD was followed for 22 days after administration and the faecal
    excretion accounted for most if not all of the elimination of TCDD
    and/or its metabolites. No radioactivity was detectable in urine and
    expired air. Twenty-two days after dosing, mean values of 1.26% and
    1.25% of the dose/g liver and adipose tissue, respectively, were
    found. Much lower 14C-activities were found in the thymus, kidney,
    and spleen, namely 0.09, 0.06, and 0.02% of the dose/g, respectively.

    The whole-body half-life was estimated to be 31 days. Rose et al.
    (1976) also followed the fate of 5 daily doses per week of 0.01, 0.1,
    and 1.0 mg 14C-TCDD/kg body weight given for 1, 3, and 7 weeks to
    male and female Sprague Dawley rats (Table 42).

         According to Kociba et al. (1976), a dose level of 0.01 g/kg per
    day, 5 days/week for 13 weeks, produced no overt toxic effects,
    whereas 0.1 or 1.0 g/kg per day produced adverse effects, including
    some deaths in the high dose group. In the study by Rose et al.
    (1976), the dose level of 0.01 g/kg per day resulted in no detectable
    14C, except in liver, fat, and excreta. Thus no kinetic calculations
    on a truly non-toxic dose could be performed. Pooling the results for
    all rats receiving 0.1 or 1.0 g/kg per day, the absorbed dose
    corresponded to 86  12% of the administered dose, with an individual
    variation of 66 to 93%. The overall rate constant for elimination
    corresponded to a half-time of 23.7 days, with an individual variation
    of 16-37 days. The radioactivity was eliminated primarily in the
    faeces, but the percentage of the dose excreted in the urine compared
    to that eliminated in the faeces tended to increase with time. At the
    dose level of 1.0 g/kg per day, males excreted in the urine 3.1 (
    0.2)% and females 12.5 ( 5.1)% of the cumulative dose over 7 weeks of
    exposure. The one female rat that died during the 7th week excreted
    17.8% of the cumulative dose in the urine. Exhaled air was not
    examined in these studies. After 7 weeks of exposure, the average body
    burdens were 47.7 ( 8.8) and 37.1 ( 7.5)%, respectively, of the
    administered dose for the rats given 0.1 and 1.0 mg TCDD/kg per day.
    The main tissue depots of 14C were the liver and adipose tissue.
    Radioactivity was also detected in thymus, kidney, and spleen at
    levels between 1 and 2% of that in liver. Direct chemical
    determination of liver samples confirmed the TCDD-concentrations
    calculated by radioactivity measurements. TCDD and/or its metabolites
    approached a steady-state body burden calculated to be 21.3 Do for
    rats given a daily dose (Do, g/kg body weight) on 5 consecutive
    days per week for an infinite number of weeks, within 13 weeks over
    the dose range 0.01 to 1.0 g/kg per day. It was thought unlikely that
    the dietary intake of extremely low levels of TCDD would result in the
    accumulation of toxic amounts in the rat.

         The excretion and distribution of a toxic dose of 400 g
    3H-TCDD/kg body weight was followed in male Sprague Dawley rats (Van
    Miller et al., 1976). Within 7 days, about 5.0% of the dose had been
    excreted in faeces and 0.5% in urine. The principal tissue depots for
    radioactivity were liver, muscle, and skin, which contained 43.0, 4.6,
    and 4.4% of the administered dose, respectively.

        Table 42. Tissue distribution of TCDD-derived radioactivity in rats given
    oral doses of 14C-TCDDa

                                                                                     

                 Tissue content of 14C (g equivalents of TCDD/kg tissue)b

                                                                                     

    Tissue             1 week              3 weeks              7 weeks
                                                                                     

                       Exposure level: 1.0 g/kg body weight per day

    liver              49.53.6            110.237.1           204.052.2
    adipose tissue     10.03.0             23.57.5             61.436.7
    thymus              0.90.2              7.34.5              6.92.5
    kidney              0.90.2              1.90.9              5.55.1
    spleen              0.40.1              1.60.8              1.91.1

                       Exposure level: 0.1 g/kg body weight per day

    liver               3.91.1             11.82.2             19.83.1
    adipose tissue      0.90.6              2.70.8              4.50.7
    thymus              ND                   1.00.6              0.60.2
    spleen              ND                   0.60.4              0.30.1
    kidney              ND                   ND                   ND

                       Exposure level:  0.01 g/kg body weight per day

    liver               ND                   0.80.1              1.60.5
    adipose tissue      ND                   0.3 (2)c             0.30.1 (4)c
    thymus              ND                   0.60.1              ND
    spleen              ND                   0.60.3              ND
    kidney              ND                   ND                   ND
                                                                                     

    a    From: Rose et al. (1976).
    b    The mean  standard deviation of 3 male and 3 female rats.
    c    Indicates the number of animals with detectable levels of
         14C-activity.
    ND   = not detected.

    
         Kociba et al. (1976) gave Sprague Dawley rats TCDD in daily doses
    of 0.01, 0.1, and 1 g/kg body weight 5 days per week for 13 weeks by
    gavage. The liver contained TCDD at a level of 324 ( 53) g/kg wet
    weight in males and 284 ( 21) g/kg wet weight in females given
    repeated TCDD doses of 1 g/kg body weight per day. For the dose level
    of 0.1 g/kg body weight per day, the liver TCDD levels were 36 ( 4)
    and 35 ( 4) g/kg wet weight for males and females, respectively. A
    dose of 0.01 g/kg body weight per day resulted in TCDD liver levels
    of 2.6 ( 0.6) g/kg wet weight in males and 3.7 ( 0.4) g/kg wet
    weight in females.

         After feeding diets with TCDD levels corresponding to daily
    dietary intakes of 0.001, 0.01, or 0.1 g/kg body weight for 2 years,
    the average concentrations of TCDD found in the liver of female
    Sprague Dawley rats were 0.54, 5.1, and 24.0 g/kg wet weight,
    respectively (Kociba et al., 1978). The corresponding levels in
    adipose tissue were 0.54, 1.7, and 8.1 g/kg wet weight. A comparison
    of liver TCDD levels found in rats given comparable daily doses of
    TCDD for 13 weeks (Kociba et al., 1976) or 2 years (Kociba et al.,
    1978) indicates that with prolonged exposure the liver TCDD content
    reaches a plateau. This finding agrees well with the prediction
    obtained by mathematical analysis of the data resulting from
    experiments involving exposure of a few weeks (Rose et al., 1976).

         The proportion of a single oral dose of 3H-TCDD found in the
    liver of female Sprague Dawley rats was dependent both on the dose
    level and on the vehicle used (Poiger & Schlatter, 1980). Maximal
    retention occurred within 48 h after dosing. Increasing retention was
    observed up to a dose of 280 ng TCDD/rat. Hepatic retention 24 h after
    dosing was higher (36.7% of the dose) if TCDD was given in 50% ethanol
    than if it was given as an aqueous suspension of soil (37%, w/w),
    where it was 16-24.1% of the dose, or activated carbon (25%, w/w),
    where it was < 0.07% of the dose (see section 7.4).

         Biliary excretion of radioactivity originating from 3H-labelled
    TCDD occurred at a more or less constant rate of 0.5 to 1% of the
    administered dose per day for 12 days following the administration of
    100 g TCDD/kg body weight in a female Sprague Dawley rat (Poiger &
    Buser, 1983). No severe toxic effects were observed within that time
    period.

         McConnell et al. (1984) dosed female Sprague Dawley rats orally
    with either pure TCDD in corn oil or amounts of contaminated soil
    giving similar doses of TCDD. In general TCDD in soil was as potent an
    inducer of aryl hydrocarbon hydroxylase (AHH) as pure TCDD in corn
    oil. The hepatic concentration of TCDD was 40.8 g/kg in the corn oil
    group, receiving 5 g TCDD/kg body weight, and 20.3 g/kg in the soil
    group, receiving 5.5 g TCDD/kg body weight (see section 7.4).

         Poiger & Schlatter (1980) studied the dermal absorption of
    3H-TCDD in hairless rats of the Naked ex Back-Cross and Holzman
    strain (200-250 g). Using the amount of TCDD-derived radioactivity
    found in the liver as an indicator of its absorption, they reported
    that the permeation of TCDD across the epidermis was highly dependent
    on the formulation used. The highest radioactivity in the liver, 14.8%
    of the administered dose, was detected when TCDD was applied as a
    methanolic solution. The hepatic recovery of the administered dose
    observed when TCDD was applied in polyethylene glycol 1500 with and
    without 15% water was 14.1 and 1.4%, respectively. Dermal application
    of TCDD in vaseline or adsorbed onto soil or activated carbon
    decreased the percentage of the dose recovered in the liver to 1.4,
    1.7-2.2, and < 0.05%, respectively.

         In studies by van den Berg et al. (1983), fly ash and crude or
    purified toluene extracts of PCDD- and PCDF-containing fly ash from a
    municipal incinerator were mixed with ordinary laboratory diet for
    rats. Small portions (2 g) of these diets were fed to male Wistar rats
    (300 g) every 24 h for 19 days, at which time the animals were
    sacrified. Tetra-, penta-, and hexa-chlorinated PCDDs and PCDFs in the
    liver and adipose tissues of these rats were determined. Rats fed the
    fly ash containing diet stored PCDDs and PCDFs in their livers at
    concentrations that were at least 3 to 5 times lower than in the case
    of rats fed comparable amounts of fly ash extracts (for the pentaCDD,
    hexaCDF, and hexaCDD isomers, the concentrations were approximately
    10-20 times lower). Generally PCDFs showed a higher retention in rat
    liver than did the corresponding PCDDs. In the adipose tissue of rats
    fed with fly ash extracts, retention was higher for penta- and
    hexaCDDs than for the corresponding PCDFs.

         In further fly ash studies, male Wistar rats (275 g) were fed for
    up to 99 days a diet that included 2.5% HC1-pretreated fly ash
    (containing PCDDs and PCDFs) from a municipal incinerator (van den
    Berg et al., 1986a). A control group received standard diet. All
    congeners retained in the liver of the rats had a 2,3,7,8-chlorine
    substitution pattern. With the exception of 2,3,4,7,8-pentaCDF and
    2,3,4,6,7,8-hexaCDF, liver retention for each congener was below 10%
    of the group dose. The retention percentages of the various congeners
    in the liver were almost equal at the time-points studied (34, 59, and
    99 days), thus indicating a long half-life of these congeners in rat
    liver.

         Male Wistar rats fed 22.7 ( 1) g or 120.7 ( 2.8) g octaCDD
    over a two-week period were found to retain about 1-2% of the given
    dose in the liver (Williams et al., 1972). The heart, kidneys, spleen,
    lung, skeletal muscle, testes, and urine contained no detectable
    levels of octaCDD, but minor amounts were found in the adipose tissue

    of the high-dose group. Faeces contained 61% and 37%, respectively, of
    the low and high dose given. The presence of a large quantity of
    octaCDD in the faeces compared to that in the bile 24-72 h after a
    single oral dose of 58 mg octaCDD to bile-cannulated male Wistar rats
    (400 g) indicated that the dioxin present in the faeces was mainly
    unabsorbed octaCDD (Williams et al., 1972).

         After 21 daily doses of 100 mg octaCDD containing 12.6 pg
    35S-thio-heptaCDD to male Sprague Dawley rats, the radioactivity was
    mainly recovered in the faeces and urine, the percentages of the
    ingested radioactive dose being 93 ( 6) and 5.2 ( 0.8)% respectively
    (Norback et al., 1975). The high faecal excretion suggests poor
    absorption. Of the radio-active body burden, 50% was contained in the
    liver. The microsomal fraction contained 96.3 ( 8.2)% of the hepatic
    radioactivity.

    6.1.2  Studies on mice

         In studies by Vinopal & Casida (1973), male white mice (20 g)
    were given tritium-labelled TCDD intraperitoneally in a single dose of
    130 g/kg body weight. Three days after TCDD administration to one
    mouse, 13% of the administered tritium was recovered in the faeces and
    0.3% in the urine, while 32% was found in the liver and 0.3% in the
    kidneys. In another study, groups of two to six mice were killed at
    various time intervals after similar treatment with the same dose of
    3H-TCDD. One and 4 days after dosing, the liver contained about 15%
    of the dose, on the 8th day, 26%, on the 11th day, 22%, and on the
    15th and 20th days, about 10%. The highest amount of 3H-activity was
    found in the microsomal fraction of the liver. Somewhat lower activity
    was detected in the mitochondrial fraction and the nuclei. The
    supernatant fraction was practically devoid of any radioactivity. On
    day 8 when the highest levels were observed, the whole liver
    homogenate contained 26.7 ( 4.8)% of the administered dose, the
    microsomes 12.6 ( 3.8)%, the nuclei 7.8 ( 1.2)%, the mitochondria
    6.2 ( 1.3)%, and the supernatant fraction 0.1 ( 0.0)% of the
    administered dose.

         Coccia et al. (1981) described the effect of adding different
    substances to food on the persistence of TCDD in the liver of male
    C57Bl/6 mice. In one of the experiments, the test diet was given
    immediately after the administration of a single oral dose of 7.6 g
    3H-TCDD/kg body weight. The hepatic radioactivity 14 days after
    dosing was 17.3, 6.3, 13.1, and 14.5% of the administered dose in
    animals fed standard chow containing 5% vegetable charcoal, 0.5%
    cholic acid, and 4% cholestyramine, respectively. When feeding of the
    test diet started 3 days after dosing, the hepatic retention of
    3H-TCDD was decreased to a similar extent.

         Faecal and urinary excretion (Table 41), along with the formation
    of faecal, biliary, and urinary metabolites (see 7.2.1.1), of
    3H-TCDD was studied in male C57BL/6, DBA/2, and B6D2F1 mice after a
    single intraperitoneal dose of 10 g 3H-TCDD/kg body weight
    (Gasiewicz et al., 1983b). The principal tissue depot in C57BL/6 and
    B6D2F1 mice was the liver, followed by the adipose tissue. Most other
    tissues examined contained less than 1% of the administered dose. In
    DBA/2 mice, the adipose tissue contained more radioactivity than the
    liver. This difference may be due to the fact that these three strains
    of mice differ in their adipose tissue content, being 5.9, 11.5, and
    5.0% of the body weight in C57BL/6, DBA/2, and B6D2F1 mice,
    respectively. The estimated half-lives of clearance of 3H-TCDD from
    the liver of C57BL/6, DBA/2, and B6D2F1 mice were 17, 27, and 13 days,
    respectively, and the corresponding figures for the half-life in the
    adipose tissue were 11, 42, and 11 days. The cumulative faecal
    elimination 30 days after dosing was 59.3, 39.8, and 56.3% of the
    administered dose in C57Bl/6, DBA/2, and B6D2F1 mice, respectively,
    and the corresponding figures for urinary elimination were 20.5, 16.8,
    and 20.5%.

         Birnbaum (1986) studied in mice the distribution and excretion of
    a single intraperitoneal dose of 500 ng (45 Ci)3H-TCDD/kg body
    weight for up to 42 days after treatment. Two sets of congenic strains
    of mice were used, i.e., male C57Bl/6 and female DBA/2 mice, where
    within each congenic pair the mice differed only at the Ah locus (or
    at a limited number of genes closely linked to the Ah locus). The mice
    were bred and phenotyped by zoxazolamine paralysis time. The results,
    some of them summarized in Tables 47 and 48, suggested that, at the
    dose level studied, the distribution and excretion of TCDD were
    primarily governed by the total genetic background rather than by the
    allele present at the Ah locus.

         When male ICR/Ha Swiss mice (27 - 35 g) were given a single oral
    dose of 135 g 14C-TCDD/kg body weight, about 71% and 1-2%,
    respectively, of the administered dose was eliminated via faeces and
    urine within the first 24 h (Koshakji et al., 1984). During the
    following 10 days, an additional 7% and 2% of the administered
    radioactivity were recovered in faeces and urine, respectively. Based
    on the estimated body burden of radioactivity, a whole-body half-life
    of 20 days was calculated.

         The distribution of 3H-TCDD in the skin of hairless (SKH:HR-1)
    mice after a single intraperitoneal dose of 6.3 g TCDD/kg body weight
    was examined for up to 14 days (Puhvel et al., 1986). Most of the
    3H-TCDD in skin was localized in the dermis, although the
    concentration of 3H-TCDD was consistently higher in the epidermis.

    6.1.3  Studies on guinea-pigs

         The retention of a single intraperitoneal dose of 2 g
    14C-TCDD/kg body weight in various tissues of male Hartley
    guinea-pigs was determined 1, 3, 5, 7, 11, and 15 days after exposure
    (Table 43) (Gasiewicz & Neal, 1979). Three animals died and all
    animals lost 24 to 35% of the body weight during the study. The
    highest amount of radioactivity per tissue was found in the liver and
    skin. The radioactivity in the liver increased with time, concomitant
    to the depletion of adipose tissues. Radioactivity in the skin
    decreased also with time. No signs of toxicity were seen when the
    cumulative excretion of a single i.p dose of 0.5 g 3H-TCDD/kg body
    weight in male Hartley guinea-pigs was studied (Gasiewicz & Neal,
    1979). The faecal and urinary excretion of radioactivity was linear
    throughout the 23-day study. Approximately 1.4% of the administered
    dose was excreted daily during that period, and the faeces contained
    94% of the excreted radioactivity.

         The microsomal fraction of the liver in male Hartley guinea-pigs
    contained 40.7 to 47.4% of the hepatic radioactivity 1 day after a
    single ip dose of 0.3, 2.0, or 7.0 g of 3H- or 14C-TCDD/kg body
    weight (Gasiewicz & Neal, 1979). Corresponding values for the crude
    nuclear fraction, the mitochondrial fraction, and the soluble fraction
    were 20.1-35.6%, 9.5-12.9%, and 7.6-26.4%, respectively. The
    subcellular distribution was similar 1 and 6 days after the low dose
    but, following the high dose, more radioactivity was present in the
    microsomal fraction and less was recovered in the crude nuclear and
    soluble fractions on day 6 after exposure.

         Olson (1986) followed the distribution, elimination and
    metabolism (see section 6.2.1.1) of a single ip dose of 0.56 g
    3H-TCDD/kg body weight in adult (335 to 625 g) male Hartley
    guinea-pigs for 45 days. One of seven animals died on day 27, but the
    remaining animals gained weight and exhibited no gross signs of
    toxicity. At termination the body composition was normal, and 61% of
    the administered radioactivity was recovered in the twelve 
    investigated tissues at the end of the study. The adipose tissue
    contained 36% of the dose; liver, pelt, and skeletal muscle plus
    carcass contained each 7% of the dose; the gastrointestinal tract
    contained about 2% of the dose, and remaining tissues contained less
    than 0.5% of the dose. Urinary and faecal elimination followed
    apparent first-order kinetics, with half-lives of 82.5 ( 22.4) and
    94.4 ( 14.7) days, respectively.



    
    Table 43. Tissue content of TCDD-derived 14C (% of dose/g tissue)a in guinea-pigs following a single
    dose of 14C-TCDDd
                                                                                                                             
                                                                Days after exposure               
                                                                                                                             
    Tissue                       1              3              5              7              11              15
                                                                                                                             
    Perirenal adipose         3.21.0        4.10.4        2.10.4        1.30.2        2.10.2
    Epididymal adipose        1.50.8        3.80.5        3.40.7        3.20.1        3.9b            2.51.1
    Adrenal                   1.40.3        1.40.2        0.90.1        1.20.3        2.10.9         1.70.2
    Liver                     1.10.4        1.50.4        1.30.2        1.10.2        2.20.2         3.20.3
    Liverc                   11.43.3       15.53.3       14.02.3       12.01.9       21.22.3        29.62.7
    Spleen                    0.70.3        0.50.3        0.20.1        0.40.2        0.40.2         0.50.1
    Duodenum                  0.40.2        0.20.1        0.20.1        0.20.1        0.20.1         0.30.1
    Pancreas                    -              -            0.20.1        0.50.3        0.40.3         0.30.1
    Stomach                   0.20.1        0.30.1        0.10.1        0.20.1        0.30.1         0.30.1
    Testes                    0.20.1        0.30.1        0.20.1        0.30.1        0.30.1         0.20.1
    Kidneys                   0.30.1        0.30.1        0.20.1        0.40.1        0.80.4         0.70.1
    Bone marrow               0.30.1        0.50.1        0.20.1        0.40.1          0.4b          0.20.1
    Lungs                     0.30.1        0.20.1        0.20.1        0.40.1        0.50.2         0.60.1
    Skinc                    13.80.7       16.30.3       15.82.4        6.50.8        6.50.7         6.70.6
    Brain, heart,
    skeletal muscle          <0.25
                                                                                                                             


    a     Mean  standard error for three animals, unless indicated otherwise.
    b     Mean of two animals.
    c     Percentage of dose/tissue.
    d     From: Gasiewicz & Neal (1979).

    


    6.1.4  Studies on hamsters

         In studies by Olson et al. (1980a), Golden Syrian hamsters
    absorbed about 73.5% of a single oral dose of 650 g 3H-TCDD/kg body
    weight, a dose that produced thymic atrophy and body weight loss in
    several of the animals. The distribution of radioactitivity in various
    tissues 1, 3, 10, and 20 days after administration is given in Table
    44. The principal depots were the liver and adipose tissue. A similar
    pattern of distribution was obtained when the same dose was given
    intraperitoneally. The elimination of radioactivity in faeces and
    urine was followed for 35 days after a single intraperitoneal (ip) or
    oral dose of 650 g/kg body weight (Olson et al., 1980a). The
    half-life for elimination was 12.0 ( 2.0) days and 15.0 ( 2.5) days
    for the ip and oral routes, respectively. Of the excreted
    radioactivity 41% occurred in urine and 59% in faeces.

         The hepatic retention of PCDDs and PCDFs from dietary intake of
    HC1-pretreated fly ash from a municipal incinerator was studied in
    male Golden syrian hamsters (van den Berg et al., 1986b). The livers
    were analysed for tetra-, penta-, and hexaCDDs and PCDFs after feeding
    the diet, which contained 25% fly ash, for 34, 58, and 95 days. No
    detectable hepatic retention was observed after 34 days. The highest
    retention after 95 days was 8.4% for 2,3,4,7,8-pentaCDF, but the
    retention was generally below 5% of the total dose. With the exception
    of 2,3,4,6,7-pentaCDF, only 2,3,7,8-substituted PCDDs and PCDFs were
    retained. Constant relative concentrations were found for the
    2,3,7,8-substituted PCDDs and PCDFs at the time points studied.

    6.1.5  Studies on monkeys

         Van Miller et al. (1976) gave three adult female rhesus monkeys
    and four male infant rhesus monkeys a single intraperitoneal dose of
    400 g 3H-TCDD/kg body weight in corn oil. This dose resulted in a
    loss of body weight: 10.8% for adults and 20.7% for infants, and light
    microscopic changes in the liver. Over a 7-day period the adult
    monkeys excreted 1.06% of the dose in the urine and 3.75% in the
    faeces. During the same period the infant monkeys excreted
    approximately 2% of the administered dose in urine and about 1.26% in
    the faeces. The authors questioned the accuracy of these figures owing
    to the difficulty of separating the two types of excreta from infant
    monkeys. The total tissue concentrations of radioactivity 7 days after
    dosing are given in Table 45. The principal tissue depots of
    radioactivity in adult monkeys were adipose tissue, skin, liver, and
    muscle; they contained 16.2, 13.1, 10.4, and 8.6% of the administered
    dose, respectively. The distribution in infant monkeys was 35.6% in
    muscle, 22.7% in skin, and only 4.5% of the dose in the liver.

    
    Table 44. Tissue distribution of TCDD-derived radioactivity in Golden
         Syrian hamsters at 1, 3, 10, and 20 days following a single oral
         dose of 650 g 3H-TCDD/kg body weighta

                                                                                     

                Tissue content of 3H (% of dose/g tissue)b

                                                                                     

    Tissue                Day 1          Day 3         Day 10         Day 20
                                                                                     

    Liver                4.031.00      5.32.82      3.19.93       0.860.09
    Liverb              12.743.21     20.443.45     9.690.99      3.700.29
    Perirenal adipose    2.930.87      3.480.56     1.380.28      0.320.03
    Adrenals             1.560.52      1.120.14     0.470.08      0.100.01
    Pancreas             0.390.20      0.610.13     0.620.26      0.210.04
    Kidneys              0.600.16      0.640.11     0.600.32      0.120.03
    Spleen               0.300.08      0.240.05     0.430.26      0.070.02
    Thymus               0.490.14      0.340.11         -          0.050.02
    Skin                 0.840.26      0.310.07     0.560.18      0.030.01
    Stomach              0.340.07      0.550.09     0.650.39      0.160.06
    Duodenum             0.510.13      0.470.09     0.550.28      0.070.02
    Jejunum              0.590.15      0.710.20     0.390.16      0.080.02
    Ileum                0.410.12      0.350.05     0.370.21      0.060.02
    Colon                0.920.27      0.600.14     0.340.07      0.060.01
    Caecum               0.390.11      0.410.10     0.280.12      0.050.01
    Lungs                0.380.09      0.370.05     0.410.25      0.070.03
    Skeletal muscle      0.200.07      0.150.05     0.150.03      0.040.02
    Heart                0.140.03      0.130.02     0.150.08      0.030.01
    Testes               0.100.04      0.320.13     0.130.04      0.030.01
    Blood                0.120.02      0.140.03     0.120.06      0.020.01
    Brain                0.030.01      0.050.01     0.060.02      0.01

                                                                                     


    a    From: Olson et al. (1980a).
    b    All values are the mean ( standard error) of four hamsters.
    c    Percentage of dose/liver.
    
    
    Table 45. Tissue distribution of TCDD-derived radioactivity in adult
         and infant rhesus monkeys 7 days following a single intraperitoneal
         dose of 400 g 3H-TCDD/kg body weighta
                                                                                  

                          (% of dose/tissue)
                          Tissue content of 3Hb
                                                 

    Tissue                Adult             Infant
                                                                               
    Liver                  10.46.9         4.511.60
    Brain                  0.580.34        1.411.40
    Spleen                0.0280.013       0.0260.004
    Small intestine        0.870.39        1.470.64
    Large intestine        1.290.12        0.640.24
    Musclec                8.622.39        35.614.4
    Skin                   13.14.9         22.78.8
    Adipose tissued        16.25.8
                                                                                  

    a     Fom: Van Miller et al. (1976).
    b     Mean ( standard deviation) of three adults or four infants.
    c     Total muscle was taken as 40% of body weight.
    d     Quantities in infant monkey were insignificant. For adult
          monkeys, the calculation was based on an estimate of 300 g
          mesenteric fat.

    
         The concentration of TCDD in samples of faeces, urine, and fat
    were measured by GC-MS at intervals after dosing up to 715 days in an
    adult female rhesus monkey (Macaca mulatta) given a single oral
    dose of 1 g TCDD/kg body weight (McNulty et al., 1982b). During the
    3 months after dosing the monkey lost 50% of its body weight but then
    began to gain weight again. The level of TCDD in faeces was high for
    4 days and then fell to very low or undetectable levels. TCDD in urine
    was very low at all time points. The apparent half-life of TCDD in
    adipose tissue was about 1 year.

         There were no significant time-dependent changes in TCDD-derived
    radioactivity found in the tissues investigated from marmosets during
    the 3-week-period following subcutaneous treatment with 5 g
    14C-TCDD/kg body weight, except for a minor decrease in levels found
    in the adipose tissue (Krowke, 1986). The data thus indicate a long
    half-life for TCDD in marmosets.

    6.1.6  Studies on dogs

         A one-year-old male beagle dog received a total dose of 5.4 g of
    TCDD enterally by direct introduction into the duodenal lumen in four
    portions of 1-2 g, with intervals of 2-7 days between treatments
    (Poiger et al., 1982). Severe toxic symtoms preceeded the death of the
    animal 17 days after the first dose. The excretion of radioactivity in
    the bile reached a maximum on day 1 or day 2 following administration.
    Significantly more biliary radioactivity was found after
    administration of doses 3 and 4 than after the administration of doses
    1 and 2, suggesting that TCDD-administration stimulated its own
    metabolism. It was later demonstrated in a 18-months-old male boxer
    that TCDD-pretreatment (10 g/kg body weight) stimulated the biliary
    excretion of 3H-labelled TCDD (32.8 ng/kg body weight), whereas
    phenobarbital-pretreatment had no effect when compared to the biliary
    excretion without pretreatment (Poiger & Schlatter, 1985). All the
    experiments were carried out in the same animal, which had enough time
    between treatments for the radioactivity to return almost to the
    background level.

    6.1.7  Studies on cows

         The major routes for elimination of 3H-TCDD in Holstein cows
    (500-650 kg) after oral doses of 0.05g (two cows) or 7.5g (one cow)
    TCDD/kg body weight were faeces > milk > urine (Jones et al., 1987).
    Fifty percent of the administered dose was eliminated in faeces, the
    major part in the first few days after treatment. Three lactating
    Holstein cows received commercial technical grade pentachlorophenol
    orally by gelatine capsule at a dose rate of 10 mg/kg body weight
    twice daily for 10 days and once daily for the following 60 days
    (Firestone et al., 1979). One cow served as a control and received
    gelatine capsules containing only ground corn. The pentachlorophenol
    composite used contained ten PCDD congeners (0.1 to 690 mg/kg) and
    eight PCDF congeners (0.9 to 130 mg/kg). Faeces collected on day 28 of
    the treatment period contained three hexaCDDs (0.05 to 0.63 g/kg),
    two heptaCDDs (21.3 to 33.1 g/kg), and octaCDD (290 to 429 g/kg).
    Faeces also contained hexa-, hepta-, and octaCDF. Milk, body fat, and
    blood contained only three of the PCDD congeners present in the
    pentachlorophenol composite, namely 1,2,3,6,7,8-hexaCDD,
    1,2,3,4,6,7,8-heptaCDD, and octaCDD. Milk samples also contained
    hexa-, hepta-, and octaCDF. The average concentrations of
    1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, octaCDD, and octaCDF in
    the composite milk fat at the end of the treatment period were 20, 40,
    25, and 2 mg/kg respectively. Similar concentrations were found in
    body (shoulder) fat at the end of the treatment period (13, 24, and 32
    mg/kg, respectively, of 1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD,
    and octaCDD). Levels of dioxins in the blood were approximately 1000
    times below the values in milk or body fat. The average daily
    excretion of 1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, and octaCDD
    in the milk during days 40-70 was about 20, 40, and 23 mg,
    corresponding to 33, 3, and 0.6% of the daily intake of PCDDs. One

    hundred days after the cessation of treatment, the average values for
    1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, and octaCDD in shoulder
    fat and milk fat were 2.5, 6.6, 5.6 mg/kg and 4.3, 6.9, 3.0 mg/kg,
    respectively.

    6.1.8  In vitro studies

         The uptake of 3H-TCDD in human fibroblasts was less efficient
    when the TCDD was associated with the high density lipoprotein (HDL)
    than with the low density lipoprotein (LDL), and even less when
    associated with serum (Shireman & Wei, 1986). From studies in mutant
    human fibroblasts lacking the normal LDL cellular receptor, the
    authors concluded that the LDL receptor pathway was involved in the
    cellular uptake of TCDD. The uptake from LDL was time, temperature,
    and concentration dependent.

    6.2  Metabolic Transformation

    6.2.1  Studies on mammals

    6.2.1.1  In vivo studies

         The possible existence of a urinary metabolite of TCDD was first
    suggested by the finding of Allen et al. (1975) that the radioactivity
    in urine of 14C-TCDD treated rats was highest from week 2 to 3. Later
    both Poiger & Schlatter (1979) and Ramsey et al. (1982) presented
    evidence for the in vivo biotransformation of TCDD in the rat.
    Male Sprague Dawley rats were given two, four, or six daily oral doses
    of approximately 15 g 14C-TCDD/kg body weight, and the bile was
    collected for 24 h following the last dose (Ramsey et al., 1982).
    Using high pressure liquid chromatography, at least eight metabolites
    of TCDD were found in the bile from these rats. Incubation of the bile
    with -glucuronidase indicated the presence of glucuronide conjugates
    among the metabolites. Poiger & Schlatter (1979) incubated similarly
    the bile or the dialysate with glucuronidase/arylsulfatase, and the
    dichloromethane-extractable radioactivity increased from 1.5% to 75%.
    Their results indicate the elimination of TCDD-metabolites in the form
    of water-soluble sulfate and glucuronide conjugates. 3H-TCDD
    metabolites extracted from the bile of one-year-old beagle dogs with
    ethanol (Poiger et al., 1982) were given to bile-duct-cannulated
    female Sprague Dawley rats (250 g) as single oral doses of 7.8-20.8 g
    3H-TCDD-metabolites/kg body weight (Weber et al., 1982b). The mean
    24-h elimination of radioactivity was 86.7 ( 6.7)% of the dose, 8.2%
    occurring in urine, 31.3% in bile, and 46.9% in faeces. A delay in the
    excretion of radioactivity in rats whose bile ducts were not
    cannulated suggested an enterohepatic circulation in the rat of the
    3H-TCDD-metabolites from the dog. The radioactive material in the
    rat bile seemed to be conjugated forms of the metabolites from the
    dogs. A metabolic breakdown scheme of TCDD in the rat and dog (Fig. 3)
    was proposed by Poiger & Buser (1983). The major metabolite seems to
    be formed via cleavage of an ether bond.

         The metabolic fate of a single ip dose of 10 g, 500 Ci TCDD/kg
    body weight in male C57Bl/6, DBA/2, and B6D2F1 mice was studied by
    Gasiewicz et al. (1983b). Samples of urine, bile, and faeces,
    collected on days 5 to 8, 8, and 7 after treatment, respectively, were
    extracted and analyzed for metabolites of TCDD by HPLC. Unmetabolized
    TCDD was detected in faeces but not in urine or in bile. More than 85%
    of the total radioactivity eliminated was present as metabolites of
    TCDD in all three mouse strains. Metabolites in the bile appear to be
    less polar than in urine. Qualitatively the elution profiles for
    urine, bile, and faeces from all three strains appeared to be quite
    similar.

         A dose of 3H-TCDD (0.56 g/kg body weight) that produced no
    gross toxicity was given ip in olive oil to six adult male Hartley
    guinea-pigs (Olson, 1986). Metabolites of TCDD were found in organic
    extracts of the liver, kidney, perirenal adipose tissue, and skeletal
    muscle in amounts corresponding to 13, 4, 8, and 28%, respectively, of
    the recovered radioactivity in these organs 45 days after dosing.
    These figures suggest that TCDD-metabolites are not efficiently
    eliminated from tissues in guinea-pigs. All radioactivity in urine and
    bile represented metabolites of TCDD, whereas in faeces most (70-90%)
    contained unchanged TCDD. Of the radioactivity administered, 73.4% was
    eliminated as unchanged TCDD in faeces and 25.7% as metabolites of
    TCDD in urine and faeces. The presence of TCDD in faeces and its
    absence in bile suggest that direct elimination of TCDD from the blood
    to the intestinal lumen may occur. The HPLC elution profiles for
    metabolites were similar, although not identical, for bile, urine, and
    tissues. Taken together, the data by Olson (1986) indicate that
    metabolism does not appear to have a major role in the ultimate
    elimination of TCDD in the guinea-pig.

         Olson et al. (1980a) collected bile and urine from Golden Syrian
    hamsters that had been treated with a single ip dose of 650 g
    14C-TCDD/kg body weight 7 days earlier. By means of HPLC, one major
    and several minor metabolites of 14C-TCDD were demonstrated both in
    bile and urine. No metabolites of 14C-TCDD were detectable in liver
    and adipose tissue, thus suggesting a rapid clearance of
    biotransformed products of TCDD.

         A one-year-old beagle dog was cholecystectomized and a Thomas
    cannula was implanted about 3 months before the first dose of TCDD
    (Poiger et al., 1982, Poiger & Buser, 1983). A total dose of 5.4 mg
    was administered enterally in four portions of 1.8, 1.08, 1.08, and
    1.44 mg on days 0, 2, 7, and 13. Five phenolic metabolites of TCDD
    excreted in dog bile were identified by combined gas
    chromatography-mass spectrometry. Severe toxic symptoms preceded the
    death of the dog 17 days after the first dose. A metabolic breakdown
    scheme of TCDD in the dog (Fig. 3) was proposed by Poiger & Buser
    (1983), lateral hydroxylation of TCDD seeming to be the major route of
    metabolism.

    FIGURE 3

    6.2.1.2  In vitro studies

         Although there is evidence of different metabolites of TCDD and
    the possibility of its metabolic transformation was suggested as early
    as 1975 (Allen et al., 1975; Rose et al., 1976), it was only in 1982
    that specific metabolites were identified by Sawahata et al. (1982).
    They incubated TCDD with isolated rat hepatocytes at 37 C for 8 h,
    and the resulting incubation mixture was subjected to HPLC. The major
    peak of radioactivity not corresponding to TCDD was incubated with
    -glucuronidase in order to split the possible glucuronide
    conjugate(s) of TCDD or its metabolite(s). They found that
    4,5-dichlorocatechol and 4,5-dichloroguaiacol are potential
    metabolites of TCDD, but due to the limited amount of material the
    identity of these metabolites was not confirmed by gas
    chromatography-mass spectrometry. Two other metabolites of TCDD,
    namely 1-hydroxy-2,3,7,8-tetrachloro-dibenzo-p-dioxin and
    8-hydroxy-2,3,7-trichlorodibenzo-p-dioxin, were isolated by means of
    HPLC, and were identified by mass spectrometry.

         Primary hepatocytes from Sprague Dawley rats and Hartley
    guinea-pigs have been used to study the metabolism of 14C-TCDD
    (Wroblewski & Olson, 1985). The overall metabolism was 2.8 times
    greater in rats than in guinea-pigs. The metabolism of 14C-TCDD was
    increased 3.2-fold in rats pretreated with TCDD (5 g TCDD/kg body
    weight ip 72 h prior to isolation of hepatocytes), but no effect was
    found in similarly TCDD-pretreated guinea-pigs, or in phenobarbital-
    pretreated rats (80 mg/kg body weight ip for 3 days, beginning 4 days
    prior to isolation of hepatocytes). Hepatocytes from TCDD-pretreated
    rats metabolized TCDD 9 times more rapidly than similarly pretreated
    guinea-pig hepatocytes. TCDD may be metabolized by an inducible form
    of cytochrome P448 which is expressed in rats but not guinea-pigs.
    These differences in metabolism may play a major role in explaining
    the differences in species susceptibility to the acute effects of
    TCDD.

    6.3  Transfer Via Placenta and/or Milk

         The transplacental passage of 14C-TCDD has been studied by
    Khera & Ruddick (1973). Pregnant Wistar rats were given 14C-TCDD in
    a single oral dose of 200 g/kg body weight on gestation days 16, 17,
    or 18 and were killed 6 h after dosing. 14C-activity was detected in
    maternal tissues and also in the fetuses and the placenta. Assuming
    that all the 14C-activity found in the samples was present as
    14C-TCDD, the following levels (ng/gram tissue) were found for
    gestation days 16, 17, and 18 respectively: maternal liver 339 ( 15),
    339 ( 19), and 275 ( 20); maternal blood 25 ( 11), 19 ( 9), and 10
    ( 3); placenta 25 ( 6), 38 ( 4), and 41 ( 3); and fetus 11 ( 3),
    15 ( 1), and 16 ( 1). Studies by Moore et al. (1973) indicated that

    the passage of TCDD or its metabolites into milk could be of
    importance, as TCDD-related effects were observed in sucklings,
    nourished by lactating mothers given, after delivery, a single oral
    dose of 1 or 3 g/kg body weight.

         The TCDD concentration in livers of pregnant NMRI mice, at a
    given dose, was significantly lower than in livers of non-pregnant
    mice (Krowke 1986). The concentration of TCDD in the liver of
    non-pregnant mice was about 5 times higher than in pregnant mice 7
    days after a s.c. dose.

         Nau & Bass (1981) studied the transfer of 14C-TCDD to embryos
    and fetuses in NMRI mice. The animals were given a single dose of 5,
    12.5, or 25 g TCDD/kg body weight by gavage or by s.c. or ip
    injection at day 16 of gestation to study the transfer of TCDD to the
    fetus. The animals were killed two days later and various tissues were
    analyzed for radioactivity. No evidence was found to indicate a major
    first pass effect following oral administration. Maternal livers
    contained the highest levels of TCDD, 4.1 to 10.5% of the radioactive
    dose administered, which was about one order of magnitude higher than
    in extrahepatic maternal tissues, including placenta. Fetal liver and
    extrahepatic tissues contained low levels of radioactivity
    corresponding to 0.09 to 1.41% and 0.05 to 0.14%, respectively, of the
    dose administered to the dams. More radioactivity was recovered in the
    placenta and fetus when TCDD was given either as a single ip dose of
    25 g/kg body weight on day 10 of gestation or as 5 daily ip doses of
    5 g/kg body weight on days 7 to 11, when compared to a single i.p.
    dose of 25 g/kg body weight on gestation day 7. Oral dosing of 30 g
    14C-TCDD (0.332 Ci/ g)/kg body weight to pregnant C57Bl/6 mice on
    gestation day 11 resulted in 0 to 14% embryomortality on gestation
    days 12 to 14 (Weber & Birnbaum, 1985). About 0.03% of the radioactive
    dose was contained in the embryo and in the placenta on days 12 to 14
    of gestation. The maternal liver contained 67.4, 67.9, and 50.6% of
    the administered radioactive dose on gestation days 12, 13, and 14,
    during which days the cumulative elimination of radioactivity in urine
    and faeces was 2.4 and 53.3% of the dose, respectively.

         The transfer of 14C-TCDD via placenta and milk and the
    distribution of the transformed TCDD between various embryonic and
    fetal tissues were studied in NMRI mice (Nau et al., 1986). Dams were
    given a single dose of 25 g 14C-TCDD (45 mCi/mmol)/kg body weight
    either orally, subcutaneously, or intraperitoneally. To differentiate
    between postnatal and in utero exposure, the experimental design
    included cross fostering. Depending on the route of administration,
    from 0.02 to 0.07% of the administered radioactivity was found in the
    liver of the fetus at birth. The highest levels were noted after ip
    administration and the lowest after oral intubation. The corresponding
    values one week after birth were 0.05 to 0.20%. The hepatic
    radioactivity in the neonate reached a peak 1 week after birth and
    then decreased slowly throughout and after lactation. The levels of

    TCDD-derived radioactivity in extra-hepatic tissues of the offspring
    were approximately one order of magnitude lower than the hepatic
    levels. Very little radioactivity was found in the stomach filled with
    milk, indicating that TCDD ingested from milk was rapidly absorbed in
    the stomach.

         Arstila et al. (1981) studied the excretion of TCDD in goat milk
    after a subchronic administration of 200 ng TCDD per day for 2 months
    in the first experiment and of 400 ng TCDD per day for one month in
    the second experiment. The minimal detectable concentration in this
    study was declared to be below 5 ng/litre. The maximum concentration
    of TCDD in milk in the first experiment was 20.8 ( 6.6) ng/litre and
    in the second experiment 19.3 ( 6.6) ng/litre. After 18 weeks feeding
    with TCDD the levels had dropped to 4.2 and 3.6 ng/litre,
    respectively.

         The secretion of TCDD in milk and cream has also been studied in
    lactating dairy cows kept on a diet containing 10, 30, 100, 300, or
    1000 mg/litre 2,4,5-T, corresponding to TCDD levels of 5, 15, 50, 150,
    and 500 ng/litre (Jensen & Hummel, 1982). This resulted in levels of
    TCDD in the excreted milk of below detection limit, 3, 10, 16-22, and
    42-89 ng/litre, respectively, indicating that about 10-20% of the dose
    given was eliminated in the milk. The levels in cream were about ten
    times higher than those in milk.

         One lactating Holstein cow, receiving commercial technical-grade
    pentachlorophenol containing several PCDDs and PCDFs orally in
    gelatine capsules at a dose rate of 10 mg/kg body weight twice daily
    for 10 days and once daily for the following 60 days, calved 151 days
    after treatment was stopped (Firestone et al., 1979). The PCDD content
    of blood, body fat (shoulder from the cow and hind quarter of the
    calf), and milk fat was determined 14 days later. The detected
    congeners were 1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, and
    octaCDD and their levels were 12, 13, 20 ng/litre and 27, 14, 6
    ng/litre in the blood of the cow and calf, respectively. Corresponding
    values for body fat were 4.8, 11.1, and 6.1 g/kg in the cow and 2.3,
    1.9, and 0.5 g/kg in the calf. The milk fat from the cow contained
    2.2, 4.4, and 3.3 g/kg of the respective congeners.

    6.4  Matrix Effects on the Uptake ("Bio-availability")

         Uptake of TCDD and other PCDD congeners is highly dependent upon
    the formulation in which it is applied. Although conflicting results
    have been obtained from studies where TCDD in soil was administered
    (McConnell et al., 1984; Umbreit et al., 1985, 1986a, 1986b), most
    available data support the idea that mixing TCDD with soil or
    activated carbon results in the adsorption of TCDD to the soil
    particles, thus reducing the availability of TCDD. Contact time of
    TCDD with soil seem to influence the availability, probably because
    the binding of TCDD to soil particles becomes strengthened (Poiger &

    Schlatter, 1980). Studies elucidating the matrix effects of various
    soils or activated carbon on the TCDD-responses in several animal
    species are summarized in Table 46.

         Poiger & Schlatter (1980) showed decreasing hepatic recovery of
    TCDD in rats 1 day after dosing using suspensions of ethanol, soil, or
    activated carbon as vehicles. About 50% lower hepatic retention of
    TCDD was obtained in the rat, 6 days after dosing, when Minker stout
    site soil was the vehicle as compared to corn oil (Lucier et al.,
    1986). However, in this study hepatic enzyme-induction (AHH and UDPGT)
    was similar in the two vehicle groups.

         The bioavailability of TCDD from environmentally contaminated
    soil samples has been studied in young guinea-pigs after intragastric
    administration (McConnell et al., 1984). Groups of six animals each
    were given soil samples corresponding to doses of approximately 1, 3,
    or 10 g TCDD per kg body weight. The doses were based on analyses of
    soil siftings (60-gauge mesh) from the Times Beach and Minker Stout
    sites, which indicated concentrations of 770 and 880 g TCDD/kg,
    respectively. Controls received soil samples in which no TCDD, PCBs,
    or PCDFs were detected. For comparison, pure TCDD in corn oil was
    given at either 0, 1, or 3 g/kg. The observation time was 30 days.
    LD50 values calculated in this study were 1.75 g/kg for TCDD in
    corn oil, 7.15 g/kg for Times Beach soil, and 5.50 g/kg for the
    Minker Stout site soil. An exact percentage for bioavailability was
    not calculated in this study, but the TCDD content of the livers of
    exposed guinea-pigs indicated a highly efficient absorption of TCDD
    from soil.

         TCDD in contaminated soil from a 2,4,5-T manufacturing plant and
    from a metal salvage yard (Newark, New Jersey, USA) had a low
    bioavailability (0.5% and 21.3%, respectively) in guinea-pigs (Umbreit
    et al., 1985, 1986a). The soils were given as single oral doses in a
    10% aqueous suspension in 5% gum acacia. The bioavailability was
    judged by the hepatic concentration of TCDD 60 days after dosing in
    guinea-pigs receiving site soils, decontaminated soil, and
    TCDD-recontaminated soil.

         The difference in bioavailability of TCDD from
    2,4,5-T-manufacturing site soil in Newark (New Jersey) and from Times
    Beach Site soil (Missouri) when given orally to guinea-pigs was
    confirmed by Umbreit et al. (1986b).

         Bonaccorsi et al. (1984) gave to rabbits seven daily doses of
    TCDD in corn oil, TCDD-contaminated Seveso soil, or recontaminated
    soil. Taking the hepatic recovery of TCDD in the corn oil group as
    100% "bioavailability", the decrease in "bioavailability" of TCDD from
    Seveso soil was 68%. The decrease in "bioavailability" of
    recontaminated soil varied from 0 to 44% with the doses used.



    
    Table 46. Matrix effects of various soils on TCDD-responses, as related to the estimated bioavailability of TCDD given
    orally to different species.
                                                                                                                             

    Species/    Vehicle/matrix          Dose of         Lethality         Hepatic           Estimated          Reference
    Length                              TCDDc                             recovery of       "bio-
    of study                                                              TCDDe             availability"
    (days)
                                                                                                                             
    Rat         50% ethanol             14.7 ng                           36.7f                                Poigner &
    1 day       recontaminated soila    12.7, 22.9 ng                     24.1f                                Schlatter
                recontaminated soilb    21.2, 22.7 ng                     16.0f                                    (1980)
                activated carbon        14.7 ng                           < 0.07f

    Rat         corn oil                1.0, 5.0                          7.6 and 40.8                         Lucier et al.
    6 days      Minker Stout site soil  1.1, 5.5                          1.8 and 20.3      50%                (1986)

    Guinea-     corn oil                1, 3            1/6, 6/6          1.6, 13.3                            McConnell et
    pig         Times Beach site soil   1.3, 3.8, 12.8  0/6, 1/5, 5/5     < 1.0-34.3        Efficient          al. (1984)
    30 days     Minker Stout site soil  1.1, 3.3, 11.0  0/6, 2/6, 6/6     < 1.0-25.7        absorption from
                recontaminated soil     10.0            6/6               45.4              soil (85%)

    Guinea-     corn oil                6               5/8                                                    Umbreit et al.
    pig         2,4,5-T-manufacturing   3, 6, 12        0/8, 0/7, 0/7     0.09 (high dose)  0.5%               (1985, 1986a)
    60 days     site soil (Newark,
                New Jersey, USA).
                Metal salvage yard soil 0.32                              0.23              21.3%
                (Newark)
                recontaminated soil     6               6/7               18
                                                                                                                             

    Table 46 (contd).
                                                                                                                             
    Species/    Vehicle/matrix          Dose of         Lethality         Hepatic           Estimated          Reference
    Length                              TCDDc                             recovery of       "bio-
    of study                                                              TCDDe             availability"
    (days)
                                                                                                                             
    Guinea-     2,4,5-T-Manufacturing   3, 5, 10        0/18, 1/20, 1/18                                       Umbreit et al.
    pig         site soil, (Newark)                                                                            (1986b)
    60 days     Times Beach site soil   1, 3, 10        2/19, 2/20, 8/14
                recontaminated soil     6               19/20

    Rabbit      corn oil                20, 40, 80d                       0.26-2.7                             Bonaccorsi et
    7 days      Seveso soil             80, 160d                          0.88-2.2          > 32%              al. (1984)
                recontaminated soil     20, 40, 80d                       0.26-1.5          > 66-100%

                                                                                                                             

    a     Contact time = 10-15 h.
    b     Contact time = 8 days.
    c     g/kg body weight, unless otherwise stated.
    d     ng/kg body weight per day.
    e     g/kg liver, unless otherwise stated.
    f     % of dose.

    


         Considerably lower hepatic levels of PCDDs and PCDFs were
    observed in rats fed a diet containing fly ash from municipal
    incinerators compared with those fed a diet containing extracts from
    the same fly ash (van den Berg et al., 1983).

         Dietary intake of soot-containing TCDD produced 60% mortality in
    male and female guinea-pigs on days 46 and 60, respectively, at which
    time the estimated TCDD-consumption was 1.3 g/kg body weight for
    males and 1.9 g/kg body weight for females (DeCaprio et al., 1983,
    see also section 8.2.3). These data thus suggest a high uptake of TCDD
    from the soot matrix.

    7.  EFFECTS OF TCDD AND OTHER PCDDs ON EXPERIMENTAL ANIMALS AND
    IN VITRO TEST SYSTEMS

    7.1  Acute Toxicity

    7.1.1  In vivo studies on mammals

         The range of doses required to cause death varies considerably
    between species, as well as between strains of species, and with sex,
    age, and route of administration within a single strain (Table 47).
    More than an 8000-fold difference exists between the dose of TCDD
    reported to cause 50% lethality to male Hartley guinea-pigs, the most
    sensitive species tested (Schwetz et al., 1973), and the corresponding
    dose for male Golden Syrian hamsters (Henck et al., 1981). The rat
    seems to be the second most sensitive species, although there is a
    more than 200-fold variability in LD50 values between different
    strains. The oral LD50 value was 22 g TCDD/kg body weight for male
    Sherman rats (Schwetz et al., 1973), whereas Walden & Schiller (1985)
    found LD50 values ranging from 164 to 340 g TCDD/kg body weight
    when male Fisher 334 N rats from three different suppliers were
    tested. The Han/ Wistar-strain of rat has been demonstrated to be
    particularly resistant to TCDD-exposure (Pohjanvirta & Tuomisto,
    1986). Among the five rats per dose group (0, 1500, 2000, 2500, or
    3000 mg TCDD/kg body weight) only one animal died within the 39-40
    days observation period.

         Monkeys (McConnell et al., 1978a), New Zealand rabbits (Schwetz
    et al., 1973), C57Bl/6 mice (Chapman & Schiller, 1985; Jones & Greig,
    1975; McConnell et al., 1978b; Smith et al., 1981; Vos et al., 1974),
    DBA/2-mice (Chapman & Schiller, 1985), and B6D2F1-mice (Chapman &
    Schiller, 1985) gave oral LD50 values of 70, 115, 114, 2570, and 296
    mg TCDD/kg body weight, respectively. The difference in sensitivity
    towards TCDD among various strains of mice has been claimed to depend
    on a genetic variability in the Ah and/or hr-locus (see section
    7.8.1).

         Male Sherman rats were found to be more sensitive to TCDD than
    were females (Schwetz et al., 1973), whereas Beatty et al. (1978)
    reported that male Sprague Dawley rats were more resistant to TCDD
    than were females. Smith et al. (1981) found adult female C57BL/10
    mice to be more resistant to TCDD than adult males of the same strain.
    No differences in sensitivity to TCDD between sexes were recorded for
    guinea-pigs (McConnell et al., 1978b; Silkworth et al., 1982) or
    hamsters (Olson et al., 1980b). Thus data on sex differences in
    sensitivity to lethal effects of TCDD are conflicting.

         Data on the effect of age at exposure to TCDD on the sensitivity
    of acute response are scarce, and comparisons are hampered by the
    absence of information, or incomplete information, on the age and/or
    body weight of the tested animals. However, Beatty et al. (1978) found

    that weanling male Sprague Dawley rats were more sensitive to TCDD
    than were adult males. A dose of 25 g TCDD/kg body weight caused,
    after 35 days, a cumulative lethality of 62% in weanling Sprague
    Dawley rats and 25% in young adults (Christian et al., 1986a). When
    weanling and mature adults were exposed to a similar toxic dose of
    TCDD (LD62 and LD60, respectively), onset of death occurred 9 days
    later in the adults (Christian et al., 1986a). 

         Schwetz et al. (1973) found LD50 values in rabbits of 115 g
    TCDD/kg body weight after oral exposure, as compared to 275 g TCDD/kg
    body weight after dermal exposure. C57Bl/6-mice seem to be more
    sensitive to ip administration of TCDD (Gasiewicz et al., 1983b) than
    to oral administration (McConnell et al., 1978b). The LD50 value in
    guinea-pigs was increased from 2.5 g/kg body weight to 19 g/kg body
    weight when the vehicle for the oral administration of TCDD was
    changed from corn oil to methyl cellulose (Silkworth et al., 1982).
    Umbreit et al. (1985) compared mortality, and time to death, among
    guinea-pigs given single oral doses of TCDD in corn oil, TCDD in a
    suspension of cleaned soil, from an industrial site, or
    TCDD-contaminated soil from the same industrial site. Animals treated
    with corn oil, cleaned soil and contaminated site soil survived the
    60-day study without any sign of TCDD intoxication. Among animals that
    received 6 g TCDD/kg body weight either in corn oil or mixed with
    cleaned soil, only 3/8 and 1/7 survived, respectively. Deaths occurred
    between days 9 to 31 and 15 to 25, respectively. These results are
    different from those reported by McConnell et al. (1984) where
    TCDD-contaminated soils from Times Beach and Minker Stout were highly
    toxic to guinea-pigs. LD50 values calculated in this study were 1.75
    g/kg for TCDD in corn oil, 7.15 g/kg for TCDD in Times Beach soil,
    and 5.50 g/kg for TCDD in Minker Stout site soil. The Minker Stout
    site soil was also potent in inducing AHH-activity in female Sprague
    Dawley rats. The different results may be due to the vehicle used
    and/or to the presence of other substances in the soils that may
    potentiate or retard the TCDD-induced toxicity (Umbreit et al.,
    1986a). In section 6.4, the results from McConnell et al. (1984) and
    Umbreit et al. (1985, 1986a) are further discussed from the viewpoint
    of the "bioavailability" of TCDD in soils (Table 46).

         Despite similar routes and vehicles for the administration of
    TCDD to Golden Syrian hamsters, LD50 values varied between 1157
    g/kg body weight (Olson et al., 1980b) and 5051 g/kg body weight
    (Henck et al., 1981). A possible explanation for this difference could
    be the spontaneous occurrence of ileitis observed in the former study,
    which might have increased the susceptibility of those hamsters to
    TCDD toxicity.



    
    Table 47. Single lethal dose values for TCDDa
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Rats

       Porton           F/5-12      8-9 weeks/     oral/DMSO           0          90 days         NR           40
                                    170-200 g                         30
    (Greig et al.,                                                    48
      1973)                                                           75
                                                                     120
                                                                     190
                                                                     300

       Porton           F/6         9-10 weeks/    oral/arachis        0          90 days         NR           40
                                    170-200 g      oil               126
    (Greig et al.,                                                   199
      1973)                                                          315
                                                                     500


       Sherman          M/5-10      NR             oral/corn oil       8          2-8 weeks       22           9-27
                                                   acetone (9:1)      16
    (Schwetz et al.,                                                  32
      1973)                                                           63

       Sherman          F/NR        NR             oral/corn oil     NR           2-8 weeks       45           13-43
                                                   acetone(9:1)
    (Schwetz et al.,
      1973)
                                                                                                                            

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Rats (contd)
       Sprague          M/6         adult/NR       ip/olive oil      4 doses      20              60           NR
       Dawley                                                        20-80        60

    (Beatty et al.,
      1978)
       Sprague          F/6         adult/NR       ip/olive oil      4 doses      20              60           NR
       Dawley                                                        10-60

    (Beatty et al.,
      1978)

       Sprague          M/6         25 days/NR     ip/olive oil      4 doses      20              25           NR
       Dawley                                                        5-50

    (Beatty et al.,
      1978)

       Fisher 334N      M/7         11-12 weeks/   oral/corn oil       0          30 days         340b         28b
                                    230-280 g                         75                          303c         26c
    (Walden &                                                        150                          164d         25d
      Shiller, 1985)                                                 225
                                                                     275
                                                                     325
                                                                     375

       CD               M/7         10-11 weeks/   oral/corn oil       0          30 days         297d         25d
                                    350-370 g                         75
    (Walden &                                                        150
      Shiller, 1985)                                                 225
                                                                     275
                                                                     325
                                                                     375
                                                                                                                           

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Rats (contd)
       Han/Wistar       M/5         NR/300-350     oral/corn oil     1500         39-40 days      > 3000       NR
                                                                     2000
    (Pohjanvirta                                                     2500
      & Tuomisto,                                                    3000
      1986)

    Mice

       C57BL/6          M/14        3 months/      oral/corn oil       0          2 months        114          15-30
                                    23.6-30.8 g    acetone (6:1)     100
    (Vos et al., 1974)                                               150
                                                                     200

       C57BL/6          M/NR        7-15 weeks/    oral/arachis      NR           35 days         126          211.6
                                    14-30 g        oil
    (Jones & Greig,
      1975)

       C57BL/6          M/8         9 weeks/       oral/corn oil     NR           30 days         284          22-25
                                    21-25 g
    (McConnell et al.,
      1978b)

       C57BL/6J         M/NR        NR             ip/olive oil      NR           30 days         132          NR

    (Gasiewicz et al.,
      1983bd)
                                                                                                                            

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Mice (contd)
       C57BL/6J         M/10-15     10-12 weeks/   oral/corn oil      95          30 days         182          24
                                    22-32 g                          145
    (Chapman &                                                       190
      Schiller, 1985)                                                285

       C57BL/10         M/5         42-121 days/NR oral/arachis       85          45 days         146          22-38
                                                   oil               107
    (Smith et al.,                                                   135
      1981)                                                          170
                                                                     213

       C57BL/10         F/5         42-121 days/NR oral/arachis       85          45 days       > 450          22-38
                                                   oil               107
    (Smith et al.,                                                   135
      1981)                                                          170
                                                                     213
                                                                     269
                                                                     338
                                                                     426
                                                                     536

       DBA/2J           M/NR        NR             ip/olive oil      NR           30 days         620          NR

    (Gasiewicz et al.,
      1983be)

       DBA/2J           M/10-15     10-12 weeks/   oral/corn oil     1370         30 days        2570          21
                                    22-32 g                          1870
    (Chapman &                                                       2610
      Schiller, 1985)                                                3500
                                                                     4470
                                                                                                                            

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Mice (contd)
       B6D2F1/J         M/NR        NR             ip/olive oil      NR           30 days         300          NR

    (Gasiewicz et al.,
      1983e

       B6D2F1J          M/10-15     10-12 weeks/   oral/corn oil     170          30 days         296          25
                                    22-32 g                          220
    (Chapman &                                                       265
      Schiller, 1985)                                                325
                                                                     425
                                                                     450

    Guinea-pigs

       Hartley          M/NR        NR             oral/corn oil     NR           2-8 weeks       0.6          5-34

    (Schwetz et al.,
      1973)

       Hartley          M/NR        NR             oral/corn oil     NR           2-8 weeks       2.1          9-42
                                                   acetone (9:1)
    (Schwetz et al.,
      1973)

       Hartley          M/6         3-4 weeks/     oral/corn oil     NR           30 days         2            17-20
                                    200-250 g
    (McConnell et al.,
      1978b)

                                                                                                                            

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Guinea-pigs (contd)
       Hartley          F/6         NR/500-600 g   oral/corn oil      0.1         42 days         2.5          32-42
                                                                      0.5
    (Silkworth et al.,                                                2.5
      1982)                                                          12.5
                                                                     20.0

       Hartley          F/6         NR/500-600 g   oral/methyl-       0.1         42 days         19           12-42
                                                   cellulose          0.5
    (Silkworth et al.,                                                2.5
      1982)                                                          12.5
                                                                     20.0

    Rabbits

       New Zealand      M,F/NR      NR             oral/corn oil     NR           2-8 weeks       115          6-39
                                                   acetone (9:1)
    (Schwetz et al.,
      1973)

       New Zealand      M,F/NR      NR             dermal/acetone     31.6        3 weeks         275          12-22
                                                                      63
    (Schwetz et al.,                                                 126
      1973)                                                          252
                                                                     500

       New Zealand      M,F/5       NR             ip/corn oil        31.6        4 weeks         NR           6-23
                                                                      63
    (Schwetz et al.,                                                 126
      1973)                                                          252
                                                                     500
                                                                                                                            

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Hamsters
       Golden Syrian    M/5-6       NR/50-80 g     ip/olive oil         0         50 days         > 3000
                                                                      500
    (Olson et al.,                                                   1000
      1980b)                                                         2000
                                                                     3000

       Golden Syrian    F/5         NR/50-80 g     ip/olive oil         0         50 days         > 3000     14-32
                                                                      500
    (Olson et al.,                                                   1000
      1980b)                                                         2000
                                                                     3000

       Golden Syrian    M/5         NR/50-80 g     oral/olive oil     500         50 days         1157         2-47
                                                                     1000
    (Olson et al.,                                                   2000
      1980b)                                                          3000

       Golden Syrian    M/6         NR/70-120 g    oral/corn oil        0         55 days         5051         9-43
                                                   acetone (9:1)      300
    (Henck et al.,                                                    600
      1981)                                                          1000
                                                                     3000
                                                                     6000
                                                                                                                            

    Table 47 (contd).
                                                                                                                            
    Species/strain      Sex/No/     Age/weight     Route/vehicle     Dose         Duration of     LD50         Time to
    (Reference)         group                                        tested       observation     (g/kg)      death
                                                                     (g/kg)                                   (days)
                                                                                                                            
    Monkeys

       Macaca           F/3         juvenile/      oral/corn oil       0          47 days         <  70       14-34
       mulatta                      2.1-2.6 kg                        70
                                                                     350
    (McConnell et al.,
      1978a)

    Dogs

       Beagle           M/2         NR             oral/corn oil      300         2-8 weeks       NA           9-15
                                                   acetone (9:1)     3000
    (Schwetz et al.,
      1973)

       Beagle           F/2         NR             oral/corn oil        30        2-8 weeks       NA           all
                                                   acetone (9:1)       100                                     animals
    (Schwetz et al.,                                                                                           sur-
      1973)                                                                                                    vived

    Chickens

       Leghorn          NR          4-6 weeks/NR   oral/NR           NR           NR              25-50        12-21

    (Grieg et al.,
       1973)
                                                                                                                            
    a     M = male, F = female, NR = not reported, NA = not applicable, ip = intraperitoneal, DMSO = dimethyl sulfoxide.
    b     supplied by Harlan.
    c     supplied by Frederick.
    d     supplied by Charles River.
    e     based on unpublished studies by Gasiewicz et al., 1983.
    


         TCDD affects a variety of organ systems in different species. The
    organ primarily affected in rodents and rabbits is the liver. In
    guinea-pigs atrophy of the thymus and lymphatic tissues seems to be
    the main effect, while dermal effects are prominent signs in non-human
    primates. Generally it is not possible to specify a single organ whose
    dysfunction is responsible for death. Overall, TCDD seems to have a
    predilection for causing pathological changes in epithelial tissues,
    both cutaneous and internal. This is particularly apparent in
    non-human primates (Macaca mulatta), and is note-worthy that the
    lesions mimic to some degree the effects in human beings. The
    histopathological alterations in tissues include hyperplastic and/or
    metaplastic alterations as well as hypoplastic responses. The toxic
    responses of various species to TCDD are summarized in Table 48,
    adapted from Poland & Knutson (1982). In all animal species studied,
    death occurred after a time lapse ranging from several days to more
    than one month after exposure. The delay was dependent on dose but not
    on species (Table 48). Progressive loss of body weight was a
    characteristic sign observed in animals given a lethal dose of TCDD.
    The weight loss became manifest usually within a few days after
    exposure and resulted in a substantial reduction of the adipose tissue
    observed at autopsy. At sublethal doses of TCDD a dose-dependent
    decrease in body weight gain occurred. This TCDD-induced wasting
    syndrome has been thoroughly investigated in several studies discussed
    more fully in section 7.4.1.

         The greatest difference between species at necropsy, both in
    gross and histological effects, concerns pathological alterations in
    the liver. As discussed in detail in section 7.4.2, a dose of TCDD
    lethal to guinea-pigs did not result in liver damage comparable to the
    liver lesions described in rabbits and rats or to liver changes
    observed in mice dying after doses higher than those needed to cause
    death in these species. In the hamster, frank liver lesions do not
    occur even after fatal doses.

         Chloracne-like lesions can be induced by topical application
    and/or systemic administration of TCDD in rabbits, non-human primates,
    and hairless mice. These lesions are further discussed in section
    7.4.4.

         Severe thymus atrophy was also found at autopsy in all animal
    species given lethal doses of TCDD. Histological examinations revealed
    lymphoid cell depletion in thymus cortex, spleen, and lymph nodes.
    These consistent findings in TCDD poisoning will be discussed in
    detail in section 7.4.5 together with the other lymphoid
    tissue-related effects.



    
    Table 48. Specific differences in toxic responses following exposure to 2,3,7,8-TCDDa, b

                                                                                                                             

                                         Monkey    Guinea-   Cowc      Rat     Mouse   Rabbitc     Chickenc       Hamster
                                                    pig
                                                                                                                             

    Hyperplasia and/or metaplasia
       Gastric mucosa                     ++          0        +        0        0                                   0
       Intestinal mucosa                   +                                                                        ++
       Urinary tract                      ++         ++       ++        0        0
       Bile duct and/or gall bladder      ++          0        +                ++                                   0
       Lung: focal alveolar                                            ++
       Skin                               ++          0        +d       0        0       ++                          0

    Hypoplasia, atrophy, or necrosis
       Thymus                              +          +        +        +        +                     +             +
       Bone marrow                         +          +                                               +
       Testicle                            +          +                 +        +

    Other responses
       Liver lesions                       +          +                ++        +       ++            +             
       Porphyria                           0          0                 +       ++                     +             0
       Oedema                              +          0                 0        +                    ++             +
                                                                                                                              

    a    References: monkey (Norback & Allen, 1973; Allen et al., 1977; McConnell et al., 1978a), guinea-pig (McConnell
         et al., 1978b; Moore et al., 1979; McConnell, 1980; Turner & Collins, 1983), cow (McConnell, 1980), rat (Kociba et
         al., 1978; Kociba et al., 1979a; McConnell, 1980), mouse (Vos et al., 1973; Schwetz et al., 1973; McConnell et al.,
         1978b), rabbit (Vos & Beems, 1971; Schwetz et al., 1973), chicken (Allen & Lalich, 1962; Vos & Koeman, 1970;
         Norback & Allen, 1973; Schwetz et al., 1973), hamster (Olson et al., 1980b; Henck et al., 1981).

    b    Symbols: 0 = lesion not observed, + = lesion observed (number of "+" denote severity),  = lesion observed to
         a very limited extent, blank = no evidence reported in literature.
    c    Responses followed exposure to 2,3,7,8-TCDD or structurally related chlorinated aromatic hydrocarbons.
    d    Skin lesions in cattle have been observed, but they differ from the skin lesions observed in other species.
    


         There are also substantial interspecies differences in the
    effects observed in other organs of animals given lethal doses of
    TCDD. Icterus was reported in rats (Buu-Hoi et al., 1972a; Gupta et
    al., 1973), hepatic porphyria occurred in mice and rats (see section
    7.4.3), and ascites with subcutaneous oedema and hydrothorax appeared
    in mice (Jones & Greig, 1975; Vos et al., 1974) and monkeys (Allen et
    al., 1977). The accumulation of serous fluid in the pericardial sac
    occurred in chickens after a single lethal dose of TCDD (see section
    7.1.4). Haemorrhages were frequently observed in many organs following
    lethal doses in monkeys (Allen et al., 1977), rats, and guinea-pigs
    (Gupta et al., 1973). In mice, death was frequently attributed to
    terminal haemorrhages (Vos et al., 1974).

          When administered in doses sufficient to cause overt toxicity,
    TCDD causes testicular atrophy and degeneration characterized by
    reduced spermatogenic activity in mice (McConnell et al., 1978b), rats
    (Kociba et al., 1976; Van Miller et al., 1977), and guinea-pigs
    (McConnell et al., 1978b). The same symptoms were present in monkeys
    fed dioxin-containing toxic fat (Allen & Carstens, 1967; Norback &
    Allen, 1973). The decreases in male sex organ weights (seminal
    vesicles, ventral prostate, testes, and caput epididymis) were
    dose-dependent. ED50 values were around 15 mg TCDD/kg body weight in
    Sprague Dawley rats 7 days after TCDD-treatment when compared to
    pair-fed control rats (Moore et al., 1985). Shrunken, hyperchromatic
    nuclei in the two layers of seminiferous tubules closest to the
    basement membrane were observed in testes of young Sprague Dawley rats
    90 h after a single ip dose of 5 mg TCDD/kg body weight (Mittler et
    al., 1984). Epididymal lesions (Khera & Ruddick, 1973) and decreased
    amounts of secretory material within the accessory sex glands (Kociba
    et al., 1976) have been reported in TCDD-treated rats. Reduced
    prostate weights, both absolute and relative, were found in Han/Wistar
    rats at non-lethal doses of TCDD (Pohjanvirta & Tuomisto, 1986).

         Reduced relative uterine weight, accompanied by decreased mucosa,
    stroma, and glands, occurred in young C57Bl/6 mice dosed with 6 g
    TCDD/kg body weight three times a week for 1 month, but there was no
    effect on the ovaries (Gallo et al., 1986). TCDD had no effect on the
    uterine weight in 25 day-old Long Evans rats 2 - 10 days after
    treatment with single ip doses of 20 or 80 g/kg body weight (Romkes
    et al., 1987). However, the increase in uterine weight induced by
    estradiol treatment was counteracted by simultaneous TCDD-treatment.

         Proliferative lesions of the gastrointestinal tract have been
    found primarily in non-human primates (Allen et al., 1977; McConnell
    et al., 1978a) whereas proliferative changes of the transitional
    epithelium in the urinary tract have been found in both guinea-pigs
    (Gupta et al., 1973; McConnell et al., 1978b) and monkeys (Norback &
    Allen, 1973).

         Reduced or unaffected spleen weight and slight to moderate loss
    of lymphocytes from spleen germinal centers have been common findings
    in laboratory animals exposed to sublethal to lethal doses of TCDD
    (Gasiewicz et al., 1980; Greig et al., 1973; Kociba et al., 1978;
    McConnell et al., 1978a; Olson et al., 1980b; Vos et al., 1973, 1974).
    Spleen cellularity in C57BL/6 mice was decreased 14 and 21 days after
    treatment with 30 g TCDD/kg body weight (Chastain & Pazdernik, 1985),
    but not in B6C3F1 or DBA/2 mice 7 days after oral doses of up to 10 g
    TCDD/kg body weight (Luster et al., 1984).

         The thyroid weight, both absolute and relative to body weight, of
    the rat has been found to be increased by TCDD-treatment (Bastomsky,
    1977; Potter et al., 1983, 1986b). Degenerative changes in the
    epithelial cells of the thyroid gland were observed in rats after 31
    daily oral doses of 1 g TCDD/kg body weight (Gupta et al., 1973) and
    7 days after a single ip dose of 150 g TCDD/kg body weight (Rozman et
    al., 1986). However, Potter et al., (1986b) found no consistent
    changes in follicle size or colloid content, variation in follicle
    size, height of follicular epithelium, or resorption of colloid at the
    periphery of follicles in rats one week after an oral dose of TCDD in
    the range 6.25 to 100 g/kg body weight.

         Histological changes in the pancreas and in the interscapular
    brown adipose tissue have been found in Sprague Dawley rats exposed to
    single ip doses of 150 g TCDD/kg body weight (Rozman et al., 1986).

         Lesions of the adrenal glands have been observed in mice and
    guinea-pigs treated with (Gupta et al., 1973; McConnell et al.,
    1978b).

         Acute exposure to lethal doses of TCDD has produced minor
    haematological alterations in all species studied. Anaemia was not
    observed in mice or guinea-pigs (Zinkl et al., 1973), but a moderate
    anaemia and leukocytosis occurred in rats (Buu-Hoi et al., 1972a) and
    monkeys (McConnell et al., 1978a). On the contrary, haemoconcentration
    was observed in rats after exposure to a somewhat lower but still
    lethal dose of TCDD (Greig et al., 1973; Zinkl et al., 1973).
    Thrombocytopenia and clotting abnormalities were observed in rats
    after acute exposure to lethal levels of TCDD (Weissberg & Zinkl,
    1973). Hypocellularity of the bone marrow was found in guinea-pigs
    (McConnell et al., 1978b), rhesus monkeys (Allen et al., 1977;
    McConnell et al., 1978a), and mice (McConnell et al., 1978b). However,
    in these studies decreased bone marrow cellularity, as judged by
    histological examination, appeared only at doses high enough to cause
    severe toxicity in the experimental animals. More recent studies
    (Chastain & Pazdernik, 1985; Luster et al., 1980, 1985) have
    demonstrated that collection and enumeration of bone marrow cells in
    suspension provides a more sensitive and quantitative method than
    histological preparation for assessing bone marrow cellularity.

    Decreased bone marrow cellularity was found in adult male C57Bl/6 mice
    3 days after exposure to 120 g TCDD/kg body weight (Chastain &
    Pazdernik, 1985), and in female B6C3F1 mice 5 days after exposure to
    10 g TCDD/kg body weight (Luster et al., 1985), but not in female
    DBA/2 mice, even at a dose of 50 g TCDD/kg body weight (Luster et
    al., 1985). The myelotoxic potential of TCDD is described in section
    7.4.6.

         Changes in clinical chemistry, including serum enzyme activities,
    serum protein concentrations, and lipid levels, observed in animals
    after acute exposure to TCDD, primarily reflect damage to other organ
    systems, mainly the liver (McConnell et al., 1978a; McConnell & Moore,
    1979; Olson et al., 1980b). Liver-related enzyme activities in serum
    are affected in those animal species where liver damage is a prominent
    sign of TCDD toxicity. In those animal species where hepatotoxicity is
    not as apparent, such as monkeys and guinea-pigs, these enzyme
    activities are essentially normal. Also the TCDD-related decrease in
    serum albumin seems to be secondary to the hepatotoxic effect, since
    the decrease is less evident or non-existent in those animals that
    show little liver damage. Generally serum triglycerides and free fatty
    acid levels are increased after TCDD exposure, while that of serum
    cholesterol is decreased. However, marked species differences exist
    and again these effects seem to be secondary to liver damage. For
    further details on hyperlipidaemia, see section 7.4.7.

         Indicators of renal damage such as blood urea nitrogen,
    creatinine, and blood electrolytes are usually within normal limits
    after TCDD exposure.

         In view of the interspecies differences in the organ distribution
    of TCDD and the variability in the effects on various organs in the
    different species, the effects of TCDD in non-human primates are of
    particular interest. Nine female juvenile rhesus monkeys (Macaca
    mulatta) were divided into three groups of three and given TCDD in
    a single oral dose of 0, 70, or 350 g per kg/body weight (McConnell
    et al., 1978a). The first indication of a toxic effect of TCDD, at day
    3, was weight loss, followed by periorbital oedema, conjunctivitis,
    and thickening of the Meibomian glands on day 12. Subsequently, the
    eye lashes, facial hair, and toe and finger nails were lost. Monkeys
    given the highest dose showed a moderate absolute lymphopenia and
    thrombocytopenia. Serum cholesterol levels dropped, while the serum
    triglyceride levels increased. Alkaline phosphatase and total
    bilirubin were normal, but glutamic oxalic transaminase and aldolase
    were increased. A decrease in the albumin fraction of the total serum
    protein was noticed. The three monkeys given TCDD in a dose of 350
    g/kg body weight died or became moribund between days 28 and 34 after
    administration. One of the monkeys given TCDD in a dose of 70 g/kg
    body weight died 14 days after administration. At autopsy body fat was
    almost completely absent in all treated monkeys. Ascites was noticed

    in two animals and all monkeys had markedly distended and thickened
    bile ducts and gall-bladders. Small focal ulcerated areas in the
    fundus of the stomach were observed in two monkeys. Microscopic
    examination showed that the Meibomian glands were dilated and filled
    with keratinaceous debris. Squamous metaplasia of the glandular
    portion, with atrophy of sebaceous cells, was present. Occasional
    scattered necrotic hepatocytes were noted in the liver on microscopic
    examination. The gastric ulcers that were found extended into the
    lamina propria.

         Schwetz et al. (1973) reported that no signs of toxicity were
    observed in four male mice and two female rats given single oral doses
    of respectively 2 and 1 g 2,7-diCDD/kg body weight (purity
    99.6-99.8%). For octaCDD (purity 98.86%), oral doses of 1 g/kg body
    weight to five female rats and 4 g/kg body weight to four male mice
    did not cause any toxic symptoms. A sample containing two different
    isomers of hexaCDD (purity > 99%, 89:11) killed 1 of 2 and 0 of 2
    male rats when given as single oral doses of 100 and 10 mg,
    respectively. The only toxic sign observed among these rats was loss
    of body weight. These studies lasted for 2 to 8 weeks.

         The toxicities of single oral doses of nine PCDDs, including
    TCDD, in C57BL/6J mice and Hartley guinea-pigs were compared by
    McConnell et al. (1978b) (Table 49). The purity of the various isomers
    tested exceeded 97%. Groups of eight male mice and of six male
    guinea-pigs were used for each dose of any of the tested compounds,
    and the animals were followed for 30 days after administration. The
    toxic effects observed after administration of the different PCDDs
    were similar, the only difference among the congeners being the amount
    needed to produce a given effect. Progressive decrease of body weight,
    more pronounced in guinea-pigs than in mice, was observed after a
    lethal dose of any of the congeners tested. Marked reduction in
    deposited adipose tissue deposits was a constant finding in animals
    given a lethal dose of any of the PCDDs. Reduction of muscle mass and
    severe dehydration were observed in guinea-pigs, and ascites,
    subcutaneous oedema, and hydrothorax in some of the treated mice.
    Decrease of thymus weight was a constant finding in both species,
    being more pronounced in mice and in the guinea-pigs that died.
    Histological examination of the thymus of the guinea-pigs that died as
    early as 5 days after administration of PCDDs revealed scattered
    necrosis of lymphocytes throughout the cortex, with concomitant
    phagocytosis by macrophages. This was even more apparent in animals
    that died 14 days after administration, in which a noticeable decrease
    in the thickness of the cortex was observed. In the animals that had
    died by day 20 it was difficult to differentiate the cortex from the
    medulla, but little evidence of necrosis was present.

        Table 49. Estimated single oral LD50 values for some PCDDsa

                                                                               
    Chlorination                  Guinea-pigs                   Mice
    of PCDDs                      (mg/kg)b                      (mg/kg)b
                                                                               
    2,8                            > 300 000                    NR
    2,3,7                             29 444                     > 3000
    2,3,7,8                                2                        284
    1,2,3,7,8                              3                        338
    1,2,4,7,8                           1125                     > 5000
    1,2,3,4,7,8                           73                        825
    1,2,3,6,7,8                       70-100c                      1250
    1,2,3,7,8,9                       60-100c                     > 440
    1,2,3,4,6,7,8                     > 600                     NR
                                                                               

    a     From: McConnell et al. (1978b).
    b     Spearman-Karber method.
    c     Estimated range due to variability in replicates.
    NR   = not reported.

    
         In guinea-pigs that survived 30 days after a lethal dose of any
    of the congeners, the thymus was reduced to one-fourth of its size in
    controls. However, thymus histology at this time was often
    comparatively normal. A reduction of the lymphoid follicles in the
    spleen and of the Peyer's patches in the intestine was observed with
    less conspicuous necrosis, which again was not evident in the 30-day
    survivors. Striking hypocellularity was found in the sternal bone
    marrow in the guinea-pigs that died, but was less obvious in
    survivors. Similar thymic and splenic changes were found in mice.
    However, bone marrow atrophy occurred only rarely in this species and
    then it was less pronounced than in guinea-pigs. In the guinea-pigs
    that died, and occasionally in survivors, a marked hyperplasia of the
    renal pelvis was observed, invariably extending into the ureter and at
    times involving the urinary bladder mucosa. Gastrointestinal
    haemorrhages and occasional microscopic dilatation of crypts in the
    glandular portion of the gastric mucosa were observed in dead
    PCDD-treated animals of both species. Retro-orbital haemorrhages with
    exophthalmus and haemorrhages with detachment of the retina were seen
    in mice that died after a lethal dose of any of the congeners tested.
    Adrenal haemorrhages and moderate atrophy of the zona glomerulosa were
    seen in guinea-pigs that died. These changes were not observed in mice
    or in surviving guinea-pigs. In guinea-pigs, primarily in animals that
    died during the observation period, changes in the spermatogenic
    epithelium were observed, with testicular tubules containing only
    spermatogonia and Sertoli cells in severely affected animals. Reduced
    spermatogenesis, necrotic spermatocytes, and spermatozoa within the

    lumen of the testicular tubules and in the epididymis, and
    multi-nucleated giant cells within the seminiferous tubules, were
    found in the mice that died but not in those that survived. On the
    other hand, liver lesions were observed with the same frequency and
    degree of involvement in all the mice given the same dose, whether
    they died or survived. Minimal liver changes were detected in
    guinea-pigs, changes largely confined to central congestion with
    occasional degeneration of hepatocytes in dead animals. Fluorescence,
    as an indication of porphyria, was found only in mice, particularly in
    the liver but also in the incisors, cranial bones, costochondral
    junction, and stifle joint. It was dose related and detectable at
    doses several times less than the LD50.

         Haemolysis and hyperproteinaemia were found in dying animals of
    both species. In mice surviving 30 days, but not in guinea-pigs, the
    blood a-globulin level was decreased, with a resultant increase in the
    albumin/globulin ratio.

         With all PCDDs for which it was possible to establish an LD50
    values were in the range 10 to 100 times lower in guinea-pigs than in
    mice. To produce the greatest toxicity the lateral positions 2,3,7,
    and 8 must be fully chlorinated. With 2,3,7-triCDD or 2,8-diCDD the
    LD50 values were in the range 1000 to 100 000 times higher than with
    TCDD. The addition of a chlorine atom at an ortho-position, e.g.,
    1,2,3,7,8-pentaCDD, resulted in only a minor reduction of toxicity. An
    additional chlorine atom further reduced the toxic potency but the
    LD50 values of 1,2,3,4,7,8-hexaCDD and 1,2,3,6,7,8-hexaCDD remained
    comparatively low. The toxicity of 1,2,3,4,6,7,8-heptaCDD was further
    reduced. A reduction in the rate of weight gain was observable in
    guinea-pigs given doses of this compound exceeding 200 g/kg body
    weight. However, no deaths were observed during the experiment even at
    doses as high as 600 g/kg body weight.

         The acute oral toxicities of soot and benzene extracts of soot,
    containing PCDDs and PCDFs from a fire in a PCB-containing transformer
    (Binghamton, New York, USA) were determined to be 410 mg/kg body
    weight and 327 mg equivalents/kg body weight, respectively, in female
    Hartley guinea-pigs (Silkworth et al., 1982). The observation period
    was 42 days. The test substances were given in 0.75% aqueous methyl
    cellulose and the doses used were 250, 500, 750, 1000, and 1250 mg
    soot/kg body weight and benzene extracts corresponding to 4, 20, 100,
    500, and 1000 mg soot/kg body weight. An extensive investigation,
    including pathology, haematology, and serum chemistry alterations, in
    groups of six male and female Hartley guinea-pigs, 42 days after
    single oral doses of 1, 10, 100, or 500 mg Binghamton soot/kg body
    weight in 0.75% methyl cellulose, was reported by Silkworth et al.
    (1982). Control animals received 500 mg activated carbon/kg body
    weight in the same vehicle. No treatment-related differences were
    observed at 1 and 10 mg soot/kg body weight. Decreased body weight
    gain occurred in both sexes at 100 and 500 mg/kg, and decreased thymus
    weight occurred at 500 mg/kg for males and at 100 and 500 mg/kg for

    females. The kidney weight was decreased only in males at 100 and 500
    mg/kg. There were no treatment-related alterations in haematological
    values. Male guinea-pigs had significantly increased serum
    triglyceride levels at 100 and 500 mg/kg and females at 500 mg/kg
    only. Elevated aspartate amino transferase (at 100 and 500 mg/kg) and
    decreased Y-glutamyltransferase (at 500 mg/kg) were observed in the
    serum of female guinea-pigs. The only clearly dose-related
    microscopical findings in soot-exposed guinea-pigs were metaplasia of
    salivary gland interlobular duct epithelium and goblet cell
    hyperplasia of pancreatic interlobular ducts. These lesions occurred
    only in males at doses of 100 or 500 mg soot/kg body weight.
    Microscopic lesions, which tended to be more frequent and/or severe in
    treatment groups than in controls, included bile duct hyperplasia,
    hepatocellular cytoplasmic inclusions, vacuolation of the adrenal
    cortex, and focal lacrimal adenitis.

    7.1.2  In vitro studies on mammalian cells

         Over 30 cell types, including primary cultures and cells from
    established and transformed cell lines derived from various tissues of
    at least six animal species, have been examined for their response to
    TCDD (Beatty et al., 1975; Knutson & Poland, 1980a; Niwa et al., 1975;
    Yang et al., 1983). The effects studied were viability, growth rate,
    and morphological alterations. No toxic effects were observed except
    in one rat hepatoma cell line in which Niwa et al. (1975) reported
    decreased viability after exposure for 72 h to TCDD at a concentration
    of 300 nmol/litre. This concentration is high when compared to the
    LD50 in rats and mice. No effect on these cells was observable at 30
    nmol/litre.

         The biochemical responses found in primary hepatocytes from rats
    exposed to 10 g TCDD/kg body weight in vivo was not present when
    the hepatocytes were exposed in vitro to TCDD at 50, 100, or 200
    nmol/litre for 48 h (Yang et al., 1983a).

    7.1.3  Studies on birds

         Chick oedema disease first gained attention in the United States
    in 1957. An extensive outbreak among chickens occurred in that year in
    Georgia (Firestone, 1973; Sanger et al., 1958; Simpson et al., 1959).
    The cause of the disease was traced to the presence in the feed of
    toxic components later identified as chlorinated dibenzo-p-dioxins.
    TCDD was identified as one of the isomers in the mixture of
    chlorinated dibenzo-p-dioxins capable of producing chick oedema
    (Flick et al., 1973).

         Clinical signs of chick oedema disease consist of dyspnea,
    reduced body weight gain, stunted growth, subcutaneous oedema, pallor,
    and sudden death. In young chickens gasping was the first noticeable
    sign, followed by a waddling gait. Gross inspection of the birds at
    autopsy revealed an increased amount of fluid in the pericardial sac

    and pale livers with a mottled and irregular granular surface. In
    advanced stages the chickens developed a distended abdomen filled with
    fluid. Endotheliosis of the vascular system was observed at
    microscopic examination, as well as pronounced proliferation of the
    endothelium of the glomerular capillaries and necrosis of the liver
    cells. Diseased chickens developed pulmonary oedema and perivascular
    lymphocyte infiltration, as well as oedema of the cardiac muscle with
    interstitial lymphocytic infiltration (Allen, 1964; Allen & Lalich,
    1962, McCune et al., 1962; Simpson et al., 1959).

         Experimentally, chick oedema has been produced with a single dose
    of 25 to 50 g TCDD/kg body weight (Greig et al., 1973). When mixtures
    of tri- and tetraCDDs were fed at dietary concentrations of 0.01
    g/kg, the chickens developed oedema and 83% of them died (Flick et
    al., 1972).

         Potency for inducing chick oedema was compared for three
    different PCDDs: TCDD (purity 91% and > 99%), hexaCDD (purity > 99%,
    two isomers), and octaCDD (purity 98.86%) (Schwetz et al., 1973). In
    the experiments, 3-day-old white Leghorn cockerels were exposed for 20
    to 21 days to one of these congeners at several dose levels. Chick
    oedema occurred in birds given oral doses of 1 or 10 g TCDD/kg per
    day, or of 10 or 100 g hexaCDD/kg per day. Chick oedema was not
    observed in chicks maintained on a diet containing 0.1 or 0.5%
    octaCDD. The weight of the Bursa of Fabricius was significantly
    decreased in 2-week-old white Leghorn cockerels decapitated 2 days
    after three daily ip doses of 10 g TCDD/kg body weight (Sawyer et
    al., 1986).

    7.1.4  Toxicity of metabolites

         Not only has identification of several in vivo metabolites been
    achieved, but the acute oral toxicity of TCDD metabolites excreted in
    the bile of dogs has been studied in the guinea-pig (Weber et al.,
    1982a). 3H-labelled TCDD was administered directly into the duodenal
    lumen of two 1-year-old Beagle dogs in four portions of 1 to 2 mg at
    time intervals of 2 to 7 days. Excretion of radioactive material from
    pooled bile samples collected daily for 4 or 7 days, and thereafter in
    pooled samples of 2 or 3 days, was performed with a method yielding
    about 50% of the total radioactivity of the bile. The extracts
    containing the metabolites were concentrated and dissolved in
    1,3-propanediol for administration. Male Pirbright guinea-pigs were
    used in a 5-week toxicity study. Five animals in each of four
    dose-groups were given a single oral dose via gastric intubation. The
    amount of TCDD metabolites was calculated by means of radioactivity
    measurements to be 0.6, 6.0, 30.0, or 60.0 g/kg body weight. Three
    animals, one control and two in the high-dose group, died within 48 h
    after administration. One death in the high-dose group was due to
    gastric perforation during dosing. The other animals which died
    exhibited histological lesions which, according to the authors, were
    due to material coextracted from bile together with the metabolites.

    The light microscopic examination of the liver, spleen, pancreas,
    thymus, kidneys, lungs, and adrenals revealed no histological changes,
    and no other toxic effects due to the metabolites of TCDD could be
    recorded in this study. The authors concluded that TCDD metabolites
    from the dog are at least 100 times less toxic to male guinea-pigs
    than TCDD itself.

         When the metabolites 2-hydroxy-3,7,8-triCDD and
    2-hydroxy-1,3,7,8-tetra CDD were given as single ip injections of 100,
    1000, or 5000 g/kg body weight to young male Wistar rats, no decrease
    in body weight gain was noted and thymic atrophy was not seen 14 days
    later (Mason & Safe, 1986). When compared to TCDD,
    1-hydroxy-3,7,8-triCDD was at least 3 orders of magnitude less
    effective in inducing hepatic AHH and EROD activities, whereas
    2-hydroxy-1,3,7,8-tetra CDD was inactive at all dose levels tested.

    7.1.5  Modulation of the acute toxicity

         Several studies have been performed which attempt to modify the
    acute toxicity of TCDD. Manara et al. (1984) studied the effect of
    activated charcoal or cholic acids in the diet on the mortality after
    60 days and mean time to death in mice, rats, and guinea-pigs exposed
    to TCDD. Dietary levels  of charcoal (2.5%), cholic acid (0.25%), and
    dehydrocholic acid (0.5%) decreased the mortality induced in C57Bl/6
    mice by  a single sc dose of 110 mg TCDD/kg body weight from 93% to
    53,  21, and 53%, respectively. The mean time to death was prolonged
    from 29 days with a normal chow diet to 35-48 days with the above
    dietary additives. In CD rats the addition of  charcoal (2.5%) and
    cholic acid (0.15%) decreased mortality from 80% to 50 and 70%
    respectively. The mean time to death was not affected. Charcoal (5%)
    added to guinea-pig chow decreased TCDD induced mortality from 64% to
    29% and reduced the mean time to death from 29 to 14 days. These
    additions to  the diet did not prevent body weight loss or liver
    enlargement in the mice or guinea-pigs (Manara et al., 1984), but 5%
    charcoal in the diet protected against TCDD-induced thymus athrophy in
    C57Bl/6 mice 14 days after a single oral dose of 10 mg TCDD/kg body
    weight (Manara et al., 1982). The addition of 5% n-hexadecane in the
    diet increased the TCDD induced mortality from 60% to 100% and the
    mean time to death, within the 50 day study, from 19.8 to 28.8 days in
    Sprague Dawley rats treated with a single ip dose of 60 mg TCDD/kg
    body weight (Rozman, 1984). N-hexadecane itself did not affect
    animal viability.

         Increased survival times have been demonstrated in mice receiving
    daily injections of triiodothyronine (T3) after TCDD treatment (Neal
    et al., 1979) and in rats thyroidectomized before TCDD treatment
    (Rozman et al., 1984). Although thyroidectomy increases the mean time
    to death, it does not prevent TCDD-related mortality in rats (Rozman
    et al., 1985a). Thyroidectomy has been demonstrated to counteract
    TCDD-induced decreases in thymus weight and to reduce the spleen

    plaque-forming cell response (Pazdernik & Rozman, 1985). A number of 
    hepatic enzyme activities were induced by TCDD to an equal magnitude
    in thyroidectomized and normal rats (Rozman et al.,  1985b). It has
    been suggested that TCDD-treatment of rats leads to hypothyroidism, a
    possible protective mechanism against TCDD toxicity (Pazdernik &
    Rozman, 1985). However, available data on changes in T3, thyroxine
    (T4), and thyroid-stimulating hormone (TSH) levels in TCDD-treated
    rats are not sufficient to state whether the animals are functionally
    hypothyroid, euthyroid, or hyperthyroid. Results presented by Potter
    et al. (1986b) (see section 8.4.9) suggest  that TCDD-treated rats
    remain essentially euthyroid and that the altered thyroid status is
    neither a major contributor to TCDD toxicity nor a key response to
    TCDD exposure.

         Daily injections of butylated hydroxyanisole protected against
    TCDD-induced lethality in female Sprague Dawley rats, whereas vitamins
    E and A, two other antioxidants, did not (Hassan et al., 1985a,b).
    Dietary selenium, if given in an optimal dose, provides partial
    protection from the lethal effects of TCDD in female Sprague Dawley
    rats (Hassan et al., 1985c). None of these treatments could counteract
    the TCDD-induced body weight loss (Hassan et al., 1985a,b,c).

    7.2  Short-term Toxicity

    7.2.1  Studies on rats

         Data from the earliest subchronic laboratory study, in which rats
    were exposed to daily and/or weekly doses of TCDD,  were reported in
    four separate papers (published simultaneously) covering general
    effects (Harris et al., 1973), pathology (Gupta et al., 1973),
    haematological and clinical chemistry changes (Zinkl et al., 1973),
    and immuno-biological effects (Vos et al., 1973). Female CD rats were
    given TCDD by  gavage in daily doses of 0.1, 1, or 10 g/kg body
    weight for 31 days. The body weight of animals exposed to the highest
    dose started to decrease within the first week of exposure and  15/16
    animals died or became moribund 17 to 31 days after administration of
    the compound began. Pathological changes were comparable to those
    observed in rats given a single lethal dose, and included severe
    thymic atrophy and liver damage, icterus, haemorrhages in various
    organs, and the depletion of lymphoid organs. Weight gain was also
    reduced at the daily dose level of 1 g/kg body weight. However, there
    were no deaths and in the animals that were killed moderate thymic
    atrophy, slight to moderate liver damage, and, in some  of the
    animals, degenerative changes in the kidneys and in the  thyroid gland
    were reported. The weight gain was not affected  and significant
    histopathological changes were not found in rats that received 0.1
    g/kg body weight per day. A decrease in thymic weight that was
    significant on day 24, was observed at the lowest exposure level
    (Gupta et al., 1973; Harris et al., 1973). Blood samples were
    collected 3, 6, 10,  13, 17, 24, and 31 days after administration of

    TCDD began. Serum enzyme activities related to liver damage began to
    increase after 10 days of exposure and remained high until death
    occurred in the 10 mg TCDD/kg per day group. This group  of animals
    also exhibited increased serum bilirubin levels commencing on day 13.
    These parameters were only slightly affected in rats receiving 1 g
    TCDD/kg per day. Thrombocytopenia occurred at all dose levels. After
    3 days of  treatment with either 1 or 10 g/kg per day, animals had
    depressed platelets counts that remained low throughout the study. In
    the low-dose rats, platelets were significantly decreased by day 17.
    No significant leukocytopenia or lymphocytopenia occurred in rats at
    any dose level. These results are in good agreement with the results
    from a more detailed haematological study on female CD rats given
    daily oral doses of 10 g TCDD/kg body weight for 10 and 14 days
    (Weissberg & Zinkl, 1973). 

         When oral doses of 0.02, 1.0, or 5.0 mg TCDD/kg body weight were 
    given weekly to groups of 10 female CD rats for 6 weeks, all the 
    animals survived (Harris et al., 1973). However, body weight gain 
    decreased in the 5.0 g/kg group during  the exposure period, and at 
    the end of this period the thymus/body weight ratio was approximately 
    50% of the ratio found in the controls. Liver damage was reported as 
    slight at this dose level. No effect on body or thymic weight and no 
    significant histopathological changes were observed in rats given 
    1 g/kg body weight or less.

         Adult male and female Sprague Dawley rats, in groups of 12, were
    given 0, 0.001, 0.01, 0.1, and 1.0 g TCDD/kg body weight by gavage,
    5 days per week for 13 weeks (Kociba et al., 1976). At the end of the
    treatment period, five rats of each sex were killed for
    histopathological examination, while the remaining animals were
    retained for post-exposure observation. Doses of 1 g TCDD/kg body
    weight per day caused five deaths in females, with three occurring
    during treatment and two after treatment, and two deaths in males,
    both occurring in the post treatment period. Decreased body weights
    and food consumption were found at the two highest dose levels both in 
    males and females. Decreased relative thymus weight and increased
    relative liver weight occurred only in the males given the highest
    dose but in both the 0.1 and 1.0 g/kg female groups. Male rats had
    significantly depressed haematological values (packed cell volume, red
    blood cell count, and  haemoglobin) in the two high-dose groups, while
    these values were normal in all female rats. Gross, as well as
    histological, examination revealed treatment-related effects only in
    the high-dose groups with some minor findings in the 0.1 g/kg  group.
    Subcutaneous oedema, decreased sizes of testes and uteri, and a
    decreased number of corpora lutea were found at necropsy. Histological
    findings were limited to lymphoid tissues, liver, and epithelial
    linings. The lymphoid tissues, including thymus, were depleted of
    lymphocytes. The liver of both male and female rats showed pleomorphic
    and multinucleated hepatocytes. Foci of necrosis, with focal
    reticuloendothelial aggregations in the areas of parenchymal cell
    necrosis, were observed. Hyperplasia of Kupffer cells and an increased
    amount of a golden-brown pigment were noted. The hepatic changes were
    more pronounced near the periphery of the lobules. Slight hyperplasia

    of bile ducts and ductular epithelium was present. The uterus was
    lined by cuboidal epithelium in the female rats. The rats given 0.01
    mg TCDD/kg did not differ from the controls in any of these
    parameters, except for a slight increase in the mean liver to body
    weight ratio.

         Goldstein et al. (1982) exposed groups of eight female Sprague
    Dawley rats to 16 weekly oral doses of 0, 0.01, 0.1, 1.0, and 10.0 g
    TCDD/kg body weight in a study (further discussed in section 8.4.3)
    primarily aimed at investigating TCDD-induced porphyria. All animals
    given the highest dose died, or were moribund after eight to twelve
    doses and were killed. A decrease in body weight gain was seen in this
    group within one week of treatment. Decreased body weight gain was
    observed also in the 1.0 g TCDD/kg per week group, but not until
    several weeks after the start of treatment. Hepatic porphyria was
    found in 7 out of 8 animals receiving weekly doses of 1.0 g TCDD/kg
    body weight, in 1 out of 8 receiving 0.1 g/kg per week, and in none
    of the animals receiving 0.01 g/kg per week or the lethal dose of
    10.0 g/kg per week. Porphyria was not reversed after six months
    recovery from a 16-week exposure to 1.0 g/kg body weight per week.

         Feeding male Wistar rats (110 g) 0, 0.2, or 1.0 g octaCDD/kg
    diet for two weeks, resulting in a total intake of 0, 22.7 ( 1.0) or
    120.7 ( 2.8) mg octaCDD, had no effect on body weight gain, feed
    consumption, or tissue weights (liver, thymus, testes, heart, and
    kidney) (Williams et al., 1972). Congestion of the liver occurred in
    the high-dose group.

         Daily doses of 100 mg octaCDD for 21 days produced no effects on
    appearance, activity, or eating habits in male Sprague Dawley rats,
    but slightly increased relative liver weight. A moderate increase in
    the hepatic smooth endoplasmic  reticulum was noted (Norback et al.,
    1975).

    7.2.2  Studies on mice

         Oral doses of 0.2, 1, 5, or 25 g TCDD/kg body weight were given
    in corn oil to male C57Bl/6 mice weekly for four weeks (Harris et al.,
    1973; Vos et al., 1973). One animal of the 25 g/kg group died after
    24 days. Significant weight loss was observed only in the high-dose
    group. Thymic atrophy, characterized by nearly complete loss of
    cortex, occurred in the 5 and 25 g TCDD/kg body weight groups.

    7.2.3  Studies on guinea-pigs

         All 10 female Hartley guinea-pigs that received weekly oral doses
    of 1 g TCDD/kg body weight died, or were killed when moribund,
    between days 24 and 32 after the first dose (Gupta et al., 1973;
    Harris et al., 1973; Vos et al., 1973). Light microscopic findings of
    moribund or dead animals revealed severe atrophy of the cortex of the
    thymus with destruction of lymphocytes. There was lymphoid cell

    depletion  in spleen and lymph nodes. Haemorrhages, mitotic figures,
    and  loss of lipid vacuoles were observed in the adrenals. Liver
    effects were restricted to diffuse single-cell necrosis, predominantly
    in the periportal area. Haemorrhages were found in the urinary bladder
    and gastrointestinal tract. The lymphocyte count was decreased at all
    doses, whereas total leukocyte values were decreased at doses > 0.04
    g/kg. Animals that received eight weekly doses of 0.008, 0.04, or 0.2
    g TCDD/kg body weight all survived. At the 0.2 g/kg dose level,
    decreased body weight gain and decreased relative thymus weight were
    observed.

         A 90-day feeding study of TCDD in male (250-370 g) and female
    (230-340 g) Hartley guinea-pigs was performed by DeCaprio et al.
    (1986) and included extensive pathology, haematology, and serum
    chemistry on surviving animals. The diets  contained 0, 2, 10, 76, or
    430 ng TCDD/kg. Animals that received the highest dose exhibited
    severe body weight loss, decreased feed consumption, and mortality.
    When 60% mortality  was reached, on day 46 for males and day 60 for
    females, the remaining animals in these groups were sacrificed. The
    estimated total TCDD consumption at that time was 1.3 and 1.9 g
    TCDD/kg body weight for males and females, respectively. No
    treatment-related mortality was observed at the 76 ng/kg dose  level,
    which corresponded to a total estimated intake of 0.44 g TCDD/kg body
    weight over the 90 days. Decreased body weight gain and increased
    relative liver weight were seen in both sexes, whereas reduced
    relative thymus weight occurred in males only. At doses of 2 and 10
    ng/kg diet, no dose-related alterations were observed. The only
    treatment-related effect on haematology and serum chemistry parameters
    was the elevation of serum triglycerides for male guinea-pigs at the
    76 ng/kg dose level. The presence of hepatocellular cytoplasmic
    inclusion bodies in female guinea-pigs was the only significant
    microscopical finding, except for thymus atrophy. Based  on this
    study, a no-observed-effect level of 0.6 ng TCDD/kg body weight per
    day in guinea-pigs was estimated.

         DeCaprio et al. (1986) also followed the body weight changes and
    mortality during and after feeding male Hartley guinea-pigs (250-360
    g) a diet containing 430 ng TCDD/kg. The diet was fed for 11, 21, or
    35 days and was then withdrawn during a 79, 69, or 55-day recovery
    period. The rate of change of weight per day was the same as that of
    animals on a control diet, after an initial weight loss of
    approximately 10% during the first 5 days. When animals were fed the
    TCDD-diet for 21 and 35 days, a significant mortality, 10% and 70%,
    respectively, was apparent. Both body weight gain and absolute body
    weight were severely depressed in surviving animals throughout the
    study. Animals destined to die generally lost more than 20% of the
    original weight, whereas less pronounced weight losses were usually
    followed by increases in body weight during the recovery period.

         A PCB-containing transformer fire at the State office building,
    Binghamton, New York, USA, resulted in contamination of the building
    with soot-like material containing various PCDDs and PCDFs (see
    section 4.5.10). This soot, mixed with diet, was fed to groups of 10
    male (250-350 g) and female (200-350 g) Hartley guinea-pigs for 90
    days (0, 0.2, 1.9, 9.3, and 46.3 mg soot/kg diet) or 32 days (231.5 mg
    soot/kg diet) (DeCaprio et al., 1983). The total intake of soot during
    the study corresponded to approximately 0.3, 3, 13, 67, and 100% of
    the LD50 dose of Binghamton soot in guinea-pigs (Silkworth et al.,
    1982) (see section 8.1.1.1). A dose-related decrease in body weight
    gain occurred at 9.3 and 46.3 mg soot/kg diet. Body weight loss and
    decreased feed intake was evident in the animals given 231.5 mg
    soot/kg diet. Three male and three female guinea-pigs given 46.3 mg
    soot/kg died. Seven animals in the highest-dose group died within 28
    to 31 days and the remaining moribund animals were killed on day 32.
    Gross pathology revealed no effects at 0.2 mg soot/kg diet. Thymic
    atrophy occurred only in males at 1.9 mg soot/kg diet, but in both
    sexes at higher doses. The relative spleen weight was significantly
    increased at 46.3 mg TCDD/kg in both sexes. Treatment-related
    microscopical findings included metaplasia of salivary gland epithelia
    (> 1.9 mg soot/kg diet), increased number of goblet cells in
    pancreatic ducts (> 46.3 mg soot/kg diet for males), focal lacrimal
    gland adenitis (> 9.3 mg soot/kg diet for males and > 46.3 mg
    soot/kg diet for females), depletion of haematopoietic cells from the
    bone marrow (> 46.3 mg soot/kg diet for females, > 231.5 mg soot/kg
    diet for males), and hepatocellular cytoplasmic inclusion bodies (9.3
    and 46.3 mg soot/kg diet in both sexes). Fatty infiltration of the
    liver, reduced thickness of thymic cortex, and degenerative changes of
    the stomach and intestine were observed only in high-dose animals.
    Haematological alterations were observed only in animals at the 46.3
    mg soot/kg diet dose level, whereas alterations in serum chemistry
    values were found also at lower levels. Toxic effects of feeding
    Binghamton soot for 90 days were similar to the effects occurring
    after acute exposure (Silkworth et al., 1982) (see section 8.1.1.1),
    but the effects were seen at a lower total dose after subchronic
    exposure than after acute dosing. The effect seen at the lowest level
    in this study was thymic atrophy at 1.9 mg soot/kg diet, which was
    equivalent to 7.8 ng TCDD/kg body weight per day. A comparison between
    the effects of feeding pure TCDD and TCDD-contaminated soot to
    guinea-pigs (DeCaprio et al., 1986) demonstrated that pure TCDD
    produced less variability of alterations and gave a steeper
    dose-response relationship for many effects. Exposure to Binghamton
    soot characteristically resulted in salivary gland duct metaplasia and
    decreased serum sodium and potassium levels (DeCaprio et al., 1983).

         When male guinea-pigs (200 g) were fed a diet containing 2.5%
    HCl-pretreated fly ash from a municipal incinerator (Zaanstad, The
    Netherlands) for up to 95 days, the animals exhibited progressive
    weight loss, hair loss, and increased relative liver weight (van den
    Berg et al., 1986b). One animal died on day 76.



    
    Table 50. Studies on long-term exposure (excluding cancer studies) to TCDD in laboratory animals
                                                                                                                             
       Species/Strain             Sex/number/     Doses tested            Treatment                Parameters
                                  groupd                                  schedule                 monitored
                                                                                                                             
       Rats
          Sprague Dawleya         M/10                    0 ng/kg         in diet                  survival
                                                          1 ng/kg         continuously
                                                          5 ng/kg         for 65 weeks
                                                         50 ng/kg
                                                        500 ng/kg
                                                       1000 ng/kg
                                                       5000 ng/kg
                                                     50 000 ng/kg
                                                    500 000 ng/kg
                                                  1 000 000 ng/kg

          Sprague Dawleyb         M and F/10      0.001 g/kg per day     in diet                  extensive
                                                  0.01 g/kg per day      continuously             histopathology,
                                                  0.1 g/kg per day       for 2 years              haematology and
                                                                                                   clinical chemistry

    Mice
       Swiss                      M/38-44         0 g/kg/week            by gavage weekly         histopathology
                                                  0.007 g/kg per week    for 1 year
                                                  0.7 g/kg per week
                                                  7.0 g/kg per week

    Monkeys
       Macaca mulattac            F/8             500 ng/kg               continuous in            extensive
                                                                          the diet for             histopathology,
                                                                          9 months                 haematology and
                                                                                                   clinical chemistry
                                                                                                                             

    a  Van Miller et al. (1977).   b  Kociba et al. (1978, 1979a,b).
    c  Allen et al. (1977).        d  M = male; F = female.
    


    7.2.4 Studies on hamsters

         No toxic effects were reported in male Golden Syrian hamsters
    (50-70 g) given a diet containing 2.5% HCl-pretreated fly ash from a
    municipal incinerator in Zaanstad, The Netherlands for up to 95 days
    (van den Berg et al., 1986b).

    7.2.5  Studies on monkeys

         A cumulative dose of 0.2 g TCDD/kg body weight, divided into
    nine oral doses given three times per week, produced no clinical
    toxicity in female rhesus monkeys (Macaca mulatta) (McNulty,
    1984). However, clearly toxic signs did occur in those monkeys that
    received cumulative doses of 1.0 and 5.0 g TCDD/kg body weight over
    the same time period. The first signs were thickening and reddening of
    the eyelids, followed by weight loss, dryness and granularity of the
    skin, and loss of hair, and in some cases anaemia, purpura, and
    bleeding from the nose and mouth. Animals that died showed squamous
    metaplasia of the sebaceous glands, mucous metaplasia of the gastric
    mucosa, hyperplasia of biliary ductal epithelium, gingivitis, and
    hypoplasia of the bone marrow. The times to death after dose were 65
    and 116 days at 5 g/kg and 130 to 211 days at 1 g/kg.

    7.3  Long-term Toxicity

         Chronic toxicity studies performed on laboratory animals exposed
    to TCDD are summarized in Table 50. Studies on carcinogenicity are
    presented in section 7.7.

    7.3.1  Studies on rats

         In studies by Van Miller et al. (1977), male Sprague Dawley rats
    were maintained in groups of 10 on diets containing 0, 1, 5, 50, 500,
    1000, 5000, 50 000, 500 000, and 1 000 000 ng TCDD/kg for 65 weeks and
    survival was monitored. At the five highest dose levels, all animals
    died before the study was completed. The first deaths in these treated
    groups occurred by weeks 31, 31, 3, 2, and 2 of treatment,
    respectively. Groups receiving 50, 500, or 1000 ng TCDD/kg in the diet
    died from acute toxic effects including severe liver necrosis, bile
    duct hyperplasia and oedema, atrophy of spleen and thymus, and
    gastrointestinal haemorrhages.

         Groups of 50 male and 50 female Sprague Dawley rats were fed
    diets providing daily doses of 0.001, 0.01, and 0.1 g TCDD/kg body
    weight for 2 years (Kociba et al., 1978, 1979a,b). Control rats, 86
    males and 86 females, received diets to which the vehicle acetone had
    been added. The dose levels corresponded to a dietary content of 22,
    208, and 2193 ng TCDD/kg feed. Increased mortality was observed in
    females given 0.1 g/kg per day, while no increased mortality was
    observed in male rats at this dose or in animals receiving doses of

    0.01 or 0.001 g/kg per day. From month 6 to the end of the study the
    mean body weights of males and females decreased at the highest dose
    and to a lesser degree in females given 0.01 g/kg per day. During the
    course of the study, subnormal body weights were occasionally also
    recorded in the low-dose group, although during the last quarter of
    the study the body weights were similar to those of the controls.

         Increased urinary coproporphyrin and uroporphyrin were noted in
    females, but not in males, given TCDD at a dose rate of 0.01 and 0.1
    g/kg per day. Analyses of blood serum collected at terminal necropsy
    revealed increased enzyme activities related to impaired liver
    function in female rats given 0.1 g TCDD/kg per day. Necropsy
    examination of the rats surviving TCDD exposure to the end of the
    study revealed that liver effects constituted the most consistent
    alteration in both males and females. Histopathological examination
    revealed multiple degenerative inflammatory and necrotic changes in
    the liver that were more extensive in females. Multinucleated
    hepatocytes and bile duct hyperplasia were also noted. Liver damage
    was dose-related and no effect was observable at the lowest dose
    studied.

    7.3.2  Studies on mice

         Weekly oral doses of 0, 0.007, 0.7, and 7.0 g TCDD/kg body
    weight for 1 year resulted in amyloidosis and dermatitis in male Swiss
    mice (Toth et al., 1979). The incidence of these lesions was 0/38,
    5/44, 10/44, and 17/43 in the control, low-, medium-, and high-dose
    groups, respectively.

    7.3.3  Studies on monkeys

         In a study by Allen & Carstens (1967) groups of four to five
    rhesus monkeys were fed diets containing 0, 0.125, 0.25, 0.5, 1.0, and
    10.0% of fat (which had been shown to be toxic to chickens) until
    death. The "toxic fat" was later demonstrated to contain various
    PCDDs, of which 65% by mass of the total PCDDs present was TCDD
    (Norback & Allen, 1973). The survival time became shorter with
    increasing doses of "toxic fat". Mean time to death was 445 days for
    the low dose and 91 days for the high dose. Decreased food consumption
    and progressive body weight loss, compared with controls, were noted.
    Both clinical and histological changes near the time of death appeared
    similar regardless of dose. The monkeys developed subcutaneous oedema,
    progressing from the eyelids and face, ascites, and hydropericardium.
    Characteristic skin changes were observed as well as anaemia,
    leukopenia, and hypoproteinaemia. The bone marrow was hypoplastic.
    Centrilobular necrosis, bile duct hyperplasia, and multinucleated
    hepatocytes were found in the liver. In more than half of the 
    animals, there was marked hypertrophy of the gastric mucosa, with
    crypts and mucin-containing cysts penetrating into the submucosa.
    Ulcerations in the fundic and pyloric regions were observed.

         Allen et al. (1977) fed eight adult female rhesus monkeys a diet
    containing 500 ng TCDD/kg for 9 months. There after, surviving animals
    were removed from the TCDD diet and were observed for another 4
    months. No control animals were included, and so data were compared
    with pre-exposure values where possible. During the first 3 months of
    exposure animals developed periorbital oedema, acne, and loss of
    facial hair and eyelashes. By 6 months these changes were quite
    prominent in six out of eight monkeys, and a decrease in haemoglobin
    haematocrit was noticed. The animals lost weight even though their
    food intake was unaltered. Two animals died within the 9-month
    exposure period and three monkeys continued to develop toxic symptoms
    and died after 3 months on a TCDD-free diet. The three surviving
    animals continued to experience periorbital oedema and loss of hair.
    The total intake of TCDD over the 9-month period was calculated to be
    2-3 g/kg body weight. Death was preceded by severe anaemia, a
    decreased white blood cell count, and severe thrombocytopenia. Autopsy
    findings included haemorrhage into a variety of organs, ascites, and
    subcutaneous oedema. Hypertrophy, dilatation, oedema, and hydropic
    degeneration of the myocardium were noted in all animals. The biliary
    ducts showed marked dilatation. Moderate hyperkeratosis of the skin,
    with cystic keratosis of the hair follicles, was noted, and
    hypocellularity of the lymphoid tissue and the bone marrow were
    observed. The hyperplastic mucous-secreting cells of the gastric
    epithelium had invaded the submucosa, and ulceration and mucinous
    cysts were common in the modified gastric mucosa. Hypertrophy and
    hyperplasia of the epithelial lining of the biliary system were
    present. The bronchial epithelium, salivary glands, bile ducts, and
    pancreatic ducts showed metaplastic changes. Death was attributed to
    complications from the severe pancytopenia. The same pattern of
    morphological changes were reported to occur in a similar study
    performed by Barsotti et al. (1979).

         Similar, though less severe, effects were observed in four adult
    female rhesus monkeys fed a diet of 50 ng TCDD/kg for 20 months
    (Schantz et al., 1978).

    7.4  Effects Detected By Special Studies

    7.4.1  Wasting syndrome

         TCDD causes a starvation-like or wasting syndrome in several
    animal species. In young animals, or following a sublethal dose in
    adults, this response is manifested as a cessation of weight gain.
    Early studies suggested that acute or chronic treatment with TCDD
    decreased food consumption, but insufficiently to account for the
    weight loss (Allen et al., 1975, 1977; Greig et al., 1973; Harris et
    al., 1973; Kociba et al., 1976; McConnell et al., 1978a,b). To
    elucidate whether malabsorption could explain the wasting syndrome,
    the transfer of a number of nutrients has been studied with everted
    intestinal sacs from TCDD-treated rats. A transient increase in the

    serosal transfer of 59Fe in Sprague Dawley rats was reported by
    Manis & Kim (1979). Absorption of glucose (Ball & Chabra, 1981; Madge,
    1977) and lipids (Shoaf & Shiller, 1981) was decreased by TCDD
    treatment. The absorption of cobalt, galactose, and proline (Manis &
    Kim, 1977) as well as of D-galactose, L-arginine, L-histidine (Madge,
    1977), and penicillin (Manis & Apap, 1979) was reported to be
    unaffected by TCDD treatment. Leucine transport was depressed in
    Sprague Dawley rats 4 h after a single oral dose of 100 g TCDD/kg
    body weight (Ball & Chabra, 1981), whereas no effect was observed in
    Fisher rats 7 days after exposure to 80 g TCDD/kg body weight
    (Schiller et al., 1982). Neal et al. (1979) found normal absorption
    and intermediary metabolism of glucose, L-alanine, and oleate in
    guinea-pigs treated with a single oral dose of 2 g TCDD/kg body
    weight. Apparently there was no generalized impairment of intestinal
    absorption. The effects reported may well be secondary to decreased
    food consumption which by itself causes structural changes in the
    intestine (Steiner et al., 1968) as well as impaired absorption of
    nutrients (Esposito et al., 1967).

         The connection between the wasting syndrome and the lethal effect
    of TCDD has been investigated in pair-feeding and forced nutrition
    studies. Courtney et al. (1978) fed TCDD- treated female Wistar rats
    a normal pelleted diet ad libitum. Supplementation with water,
    electrolyte solution, or liquid diet, administered by gavage, could
    not reverse or change the pattern or extent of TCDD-induced weight
    loss or mortality.

         To bypass gastrointestinal absorption, Gasiewicz et al. (1980)
    fed rats intravenously with total parenteral nutrition (TPN). Rats
    that had received a single ip dose of 100 g TCDD/kg body weight
    gained weight similarly to their TPN-fed controls, yet still died at
    days 13 to 17 following treatment. TCDD-treated rats fed a chow diet
    ad libitum lost weight progressively (as compared to pair-fed
    controls that maintained their starting weight) and died at days 11 to
    20. In TPN-fed TCDD-treated rats, liver damage was more severe and fat
    depots were increased as compared to chow-fed TCDD-treated animals.
    Similar results were obtained with TPN-fed male Hartley guinea-pigs
    treated with a single ip dose of 2 g TCDD/kg body weight in olive
    oil, when compared to TPN-fed control guinea-pigs (Lu et al., 1986).
    Similar signs of toxicity were present in TPN- and ad libitum-fed
    TCDD-treated guinea-pigs. In contrast to TPN-fed rats (Gasiewicz et
    al., 1980), TCDD- treatment in TPN-fed guinea-pigs only produced mild
    hepatic changes, including increased liver lipid and cytochrome P 450
    content, but no morphological changes (Lu et al., 1986).

         Seefeld et al. (1984a) suggested that TPN-fed TCDD- treated rats
    might have suffered from overnutrition and, secondary to that,
    enhanced hepatotoxicity, as compared to chow-fed, TCDD-treated rats.
    These same investigators have presented a heuristic model for the

    TCDD-induced wasting syndrome based on the assumption that body weight
    in rats is regulated around an internal standard or set point (Keesey
    et al., 1976). Prevailing weight at a given age is constantly being
    compared to this set point value and if differences occur, feed
    consumption is adjusted so as to raise or lower body weight to match
    the set point level. If TCDD lowers this setpoint, reduction in food
    consumption would result, as the rat attempts to reduce its weight to
    a new lower level of regulation determined by the dose of TCDD
    administered. This hypothesis has been tested in several experiments
    under carefully controlled feeding procedures.

         Repeated studies have demonstrated that reduction of feed intake,
    due to increased food spillage, is sufficient to account for the loss
    of body weight in TCDD-treated Sprague- Dawley rats (Seefeld &
    Peterson, 1983, 1984; Seefeld et al., 1984a,b). TCDD-treated rats
    maintain and defend their reduced weight level with the same precision
    as control rats (fed ad libitum) defend their normal weight level
    (Seefeld et al., 1984b). The percentage of the daily feed intake that
    is absorbed by the gastrointestinal tract of TCDD-treated and control
    rats is similar (Potter et al., 1986a; Seefeld & Peterson, 1984).
    Water intake, resting and total oxygen consumption, carbon dioxide
    production, respiratory quotient, and spontaneous motoractivity were
    decreased in a dose-dependent manner by TCDD-treatment (Potter et al.,
    1986a; Seefeld & Peterson, 1983; Seefeld et al., 1984a). Urine output
    was unaffected by TCDD-treatment, despite decreased water intake,
    whereas urinary excretion of energy and urea were decreased and
    urinary ammonia was increased (Potter et al., 1986a).

         Hypophagia was the major cause of the loss of adipose and lean
    tissue in male Fisher F-344 rats, C57Bl/6 mice, and albino guinea-pigs
    when exposed to a calculated LD80 dose of TCDD (Kelling et al., 1985).
    Body weight loss followed a similar time course in TCDD-treated and
    pair-fed control animals of all three species. Lethalities for
    TCDD-exposed rats, mice, and guinea-pigs were 95%, 69%, and 81%,
    respectively, compared to lethalities in the appropriate pair-fed
    controls of 48, 14, and 64%, respectively (Kelling et al., 1985).
         Lethality and body weight loss followed almost identical
    time-course-curves in Sprague Dawley rats that received a single oral
    dose of 75 g TCDD/kg body weight and in pair-fed controls (Christian
    et al., 1986a). Thus the contribution to lethality made by body weight
    loss seems to depend on the species and strain. Weight loss appears to
    play a greater role in causing death in Sprague Dawley rats and
    guinea-pigs than in Fisher F-344 rats and C57Bl/6 mice. Christian et
    al. (1986a) demonstrated differences in organ weights and
    histopathology in TCDD-treated and pair-fed animals, despite similar
    time-courses and magnitude of body weight loss and lethality. Pair-fed
    animals exhibited lesions in the gastro-intestinal tract, which were
    absent in TCDD-treated rats, that may have contributed to death. The
    hepatic carbohydrate, protein, and lipid metabolism was affected

    differently in TCDD-treated and pair-fed Sprague Dawley rats
    (Christian et al., 1986b). To distinguish direct effects of TCDD from
    effects secondary to hypophagia, the studies by Christian et al.
    (1986a,b) and Potter et al. (1986a) were performed with schedule-fed
    animals. The reason for this was the finding that the 24-h feeding
    pattern for TCDD-treated rats was different from the feeding pattern
    for pair-fed controls, although it was similar to that of control rats
    fed ad libitum. Decreased feed consumption did not contribute to
    weight loss in C57BL/6 mice exposed to TCDD until the animals were
    moribund (Chapman & Schiller, 1985).

         Besides being typical signs of TCDD toxicity, loss of body weight
    and appetite are also prominent signs of thyroid dysfunction (see
    section 8.4.9). Serum glucose levels were also decreased by TCDD
    independently of hypophagia, whereas the decrease in serum insulin
    appeared to result from hypophagia, since it was seen in both
    TCDD-treated and pair-fed controls. These results indicate that the
    effect of TCDD on thyroid hormones cannot explain the TCDD-induced
    decrease in body weight gain.

         An interesting biochemical effect in TCDD-induced wasting is the
    ability of TCDD to decrease hepatic vitamin A storage in rats (see
    section 8.4.10). It has long been known that vitamin A is necessary
    for growth and that vitamin A deficiency will result in depressed body
    weight gain as well as in reduced food intake. However, the animal
    continues to eat and grow though body weight gain is less than normal
    (Brown & Morgan, 1948; Coward, 1947; Hayes, 1971; Orr & Richards,
    1934; Patterson et al., 1942).

         The effect of chemical structure on the ability of several PCDDs
    to cause body weight loss in rats has been investigated (Mason et al.,
    1986). Of the congeners studied, 2,3,7,8-TCDD was the most active.
    Those congeners fully substituted in the 2,3,7, and 8 positions but
    containing additional chlorosubstituents in the non-lateral 1,4,6, and
    9 positions were less active.

    7.4.2  Hepatotoxicity

         TCDD produces hepatomegaly, due to hyperplasia and hypertrophy of
    parenchymal cells, in all species that have been investigated, even at
    sublethal doses. However, there is considerable variation between
    species in the extent of this lesion. Other liver lesions are more
    species specific.

         Liver lesions alone cannot explain lethality following TCDD
    administration, though it may be a contributing factor at least in the
    rat and rabbit.

         The morphological changes in the liver are accompanied by
    impaired liver function, characterized by liver enzyme leakage,
    increased microsomal monooxygenase activities, porphyria, impaired
    plasma membrane function, hyperlipidaemia, and increased regenerative
    DNA synthesis.

    7.4.2.1  Morphological alterations

         In Charles River rats given single oral sublethal doses of TCDD
    (5 or 25 g/kg body weight) a dose-related increase was observed 3
    days after dosing in the amount of hepatic smooth endoplasmatic
    reticulum (SER) around the periphery of cells, particularly in the
    areas around bile canaliculi. The effect progressed by days 6 and 9,
    when an increased amount of rough endoplasmatic reticulum (RER) was
    also present. By day 28 these changes had returned essentially to
    normal levels (Fowler et al., 1973).

         The livers of CD rats given high sublethal doses of TCDD (5.0
    mg/kg body weight per week) for six weeks showed transient
    degenerative changes, followed by megalocytosis, regeneration, and the
    occurrence of multinucleated giant hepatocytes (Gupta et al., 1973).
    They also showed that the hepatotoxic reaction in rats given lethal
    doses of TCDD (10 g/kg body weight per day for 16-31 days) was
    characterized by degenerative and necrotic changes, with the
    appearance of mononuclear cell infiltration, multinucleated giant
    hepatocytes, increased numbers of mitotic figures, and pleomorphism of
    cord cells. These lesions were considered severe enough to be a
    contributing factor to death.

         Parenchymal cell necrosis was observed by Greig et al. (1973) in
    Porton rats 3 weeks after exposure to an LD50 dose of TCDD. The
    necrosis, which was located in the centrilobular zone close to the
    central vein, became more severe with time.

         Jones & Butler (1974) further investigated the time course of the
    TCDD-induced liver lesions appearing in the centrilobular zone. They
    confirmed the transient degenerative and inflammatory lesions
    previously reported (Greig et al., 1973; Gupta et al., 1973). At the
    ultrastructural level, consistent changes occurred in the cytoplasm
    whereas normal nuclear morphology and division were found throughout
    the study. Two weeks after a single oral dose of 200 g TCDD/kg body
    weight to Porton rats, extensive fusion of parenchymal cell plasma
    membranes in the centrilobular zone was replaced by a diffuse zone
    with islands of normal membrane occurring at intervals. Normal tight
    and gap junctions were present in control animals and in periportal
    areas of the test animals. These findings suggest that the
    multinucleated cells occurring in TCDD-treated rats might form by
    coalescence of parenchymal cells. The effect of TCDD on plasma

    membranes demonstrates a specific subcellular site of action, which
    might be involved in the toxic action of TCDD. This lesion was,
    however, not observed until 2 weeks after treatment and thus could not
    explain the immediate effects on, for instance, food intake, body
    weight gain, and general health.

         The time course for liver lesions in male Sprague Dawley rats
    (200 g) given a single ip dose of 20 g TCDD/kg body weight was
    followed for up to 32 weeks by Weber et al. (1983). The lesions became
    progressively worse up to the 16th week after injection, and
    thereafter appeared to regress slowly. The lesions were almost
    identical to those reported previously (Fowler et al., 1973; Jones &
    Butler, 1974). The histological findings were accompanied by
    hyperbilirubinaemia, hypercholesterolaemia, hyperproteinaemia, and
    increased serum glutamicoxaloacetic transaminase and serum
    glutamic-pyruvic transaminase activities, further indicating decreased
    liver function (Greig et al., 1973; Zinkl et al., 1973).

         Fewer studies of liver lesions produced by TCDD have been made in
    other species than in the rat.

         In C57BL/6 mice given single oral doses of 100, 150, or 200 g
    TCDD/kg body weight (Vos et al., 1974) and 250 g TCDD/kg body weight
    (Jones & Greig, 1975), centrilobular degenerative and necrotic changes
    were present but multinucleated parenchymal cells were not seen.
    Proliferation of the bile ducts and bile duct epithelial cells, as
    well as lipid accumulation, have been observed, with a substantial
    increase in the hepatic levels of esterified fatty acids and
    cholesterol. Only slight damage, hepatocellular swelling, was reported
    in CD-1 mice 21 days after a single dose of 50 g TCDD/kg body weight,
    and no histological changes were detected 7 or 35 days after
    administration (Gupta et al., 1973).

         The guinea-pig, while being very sensitive to TCDD, as indicated
    by LD50 data, shows less severe morphological alterations in the
    liver than do other species. No manifest liver lesions at the light
    microscopical or ultrastructural levels has been found (Gupta et al.,
    1973; McConnell et al., 1978b; Moore et al., 1979; Turner & Collins,
    1983).

         The hamster is very resistant to TCDD toxicity (Table 47) and
    exhibits no manifest liver damage even after a fatal dose (Henck et
    al., 1981; Olson et al., 1980b). However, Gasiewicz et al. (1986)
    found bile duct hyperplasia, numerous inflammatory cells, and
    increased number of multinucleated cells in the livers of male Golden
    Syrian hamsters 35 days after they received a single ip dose of 500 g
    TCDD/kg body weight in olive oil.

    7.4.2.2  Hepatic plasma membrane function

         The morphological impairment of hepatic plasma membranes in the
    centrilobular parenchymal cells of TCDD-treated Porton rats was
    demonstrated histochemically to be preceeded by a loss of ATPase
    activity (Jones, 1975). The effect occurred 3 days after treatment in
    an area five to six cells deep around the central vein along the
    canalicular borders, and became more severe with time. At the end of
    the study (day 42 posttreatment) the ATPase activity was completely
    abolished around the central vein, including the mid-zonal region, and
    encroached on the periportal area in moribund animals. The loss of
    ATPase activity was related to the clinical state of the animal. Thus
    animals displaying minimal signs of intoxication retained the normal
    distribution of ATPase in the periportal zone. In animals killed 9
    months after treatment, partial restoration of the normal liver
    architecture and the ATPase activity were evident.

         Biochemical studies of isolated heptatic plasma membranes from
    Holtzman rats treated with 10 or 25 g TCDD/kg body weight revealed
    depressed ATPase activities (Peterson et al., 1979a). The activity of
    Na/K-ATPase was depressed to the same extent for both doses from day
    2 to 40 after treatment, while a similar depression of Mg-ATPase
    activity was observed only in the high-dose group. The Mg-ATPase
    activity tended to recover by day 40, whereas Na/K-ATPase activities
    did not. A pair-feeding experiment demonstrated that these effects
    were independent of the TCDD-induced decrease in food consumption. In
    vitro incubation of plasma membranes indicated that ATPase inhibition
    did not occur by direct interaction with TCDD. Greig & Osborne (1981)
    demonstrated a decrease in K/Mg-ATPase activity, but not in the
    Na/K-ATPase or 5'-nucleotidase activities of hepatic plasma membranes
    prepared from female Porton rats 6 and 11 days after a dose of 200 g
    TCDD/kg body weight.

         Many physiological homeostatic mechanisms are dependent on proper
    plasma membrane function and composition. Matsumura et al. (1984)
    reported that a single ip dose of 25 g TCDD/kg body weight to male
    Sprague Dawley rats had reduced hepatic plasma membrane
    ATPase-activities by 40% 10 days after treatment. The marker enzyme
    for putative preneoplastic hepatocytes, glutamyl transpeptidase, was
    reduced, while protein kinase (both c-AMP-stimulated and
    c-AMP-nonstimulated) was increased (Matsumura et al., 1984). Both
    c-AMP-dependent and c-AMP-independent protein kinase, in hepatic
    plasma membranes from male Sprague Dawley rats treated with an ip dose
    of 25 g TCDD/kg body weight, were maximally increased on day 20 after
    treatment (4.5- and 12-fold, respectively). The induction was
    measurable within 2 days after the administration and was still
    persistent after 40 days (Bombick et al 1985). Protein kinase C was
    significantly increased in hepatic plasma membranes from Sprague
    Dawley rats but not from guinea-pigs 10 days after single ip doses of
    25 and 1 g TCDD/kg body weight, respectively (Bombick et al., 1985).
    TCDD treatment in vivo also affected the in vitro binding of

    concanavalin A, epidermal growth factor, and insulin to their cell
    surface membrane receptors.

         The binding of glucagon and prostaglandin E were not affected by
    a dose of 25 g TCDD/kg body weight (Matsumura et al., 1984). Studies
    with Sprague Dawley rats demonstrated that the TCDD-induced decrease
    in epidermal growth factor binding (EGF) was observable within 2 days
    after dosing, reached its maximum by day 20, and was still significant
    40 days after dosing. The decrease was observable after a single ip
    dose of 0.1 g TCDD/kg body weight (Madhukar et al., 1984). The
    relative doses of TCDD needed to suppress EGF binding to 50% of the
    control level were 1, 14, and 32 g/kg body weight for the guinea-pig,
    the Sprague Dawley rat, and the Golden Syrian hamster, respectively
    (Madhukar et al., 1984). A single ip dose of 115 g TCDD/kg body
    weight had decreased the EGF binding 10 days after treatment by 93.1,
    97.8, and 46.0% in C57Bl/6, CBA, and AKR mice, respectively (Madhukar
    et al., 1984). The effect of daily sc injections of 2 ng EGF/kg body
    weight to newborn Balb/c-mice was compared with the effect of a single
    ip dose of 10 mg TCDD/kg body weight given to the dam of newborn
    Balb/c-mice within 3 h of delivery (Madhukar et al., 1984). The
    parameters studied, all well known in vivo effects of EGF, were:
    time for eyelid opening and tooth eruption; hair length and diameter
    on day 14; and body weight and thymus weight on day 22. All parameters
    were significantly reduced both by TCDD treatment and EGF treatment,
    as compared to controls.

         Bombick et al. (1984) found that, 10 days after treatment, in
    vitro binding of 125I-low density lipoprotein (LDL) to its receptor
    on hepatic plasma membranes was decreased by 73% in TCDD-treated
    guinea-pigs (1 g TCDD/kg body weight) as compared to pair-fed
    controls. Primary hepatocytes from guinea-pigs treated in the same way
    had a reduced ability to internalize 125I-LDL. The reduction of LDL
    receptors on hepatic plasma membranes might be responsible for the
    increase in plasma of very low density lipoproteins (VLDL) and LDL
    noted in TCDD-treated animals.

         Quantitative changes in the protein composition of plasma
    membranes following TCDD treatment have been reported (Brewster et
    al., 1982; Matsumara et al., 1984). The membranes were isolated from
    male Sprague Dawley rats 10 days after an ip injection of 25 g
    TCDD/kg body weight and analyzed by SDS-polyacrylamide gel
    electrophoresis. Some small proteins (14 000-30 000 daltons) were
    completely abolished by TCDD treatment. The effect was most pronounced
    10 to 20 days after treatment.

    7.4.2.3  Biliary excretion

         The early proliferation of liver cells around bile canaliculi
    seen after TCDD treatment (Fowler et al., 1973) was suggestive of an
    effect on biliary excretion.

         The cumulative biliary excretion of indocyanine green (ICG), an
    organic anion, was decreased in a dose-dependent manner by treatment
    with 5 or 25 g TCDD/kg body weight in CD rats (Hwang, 1973). On the
    contrary, biliary excretion of the organic anions sulfobromophthalein
    and phenol-3,6-dibromophthalein was unaffected by treatment with 10 or
    25 g TCDD/kg body weight in Holtzman rats (Yang et al., 1977).

         Biliary excretion of ouabain, a model compound for neutral
    non-metabolized substrates such as estradiol, progesterone, and
    cortisol, was depressed in male Holtzman rats in a dose-related manner
    by a single oral dose of 10 or 25 g TCDD/kg body weight (Yang et al.,
    1977). The effect was detectable two days after treatment, reached a
    peak between 10 and 20 days, and recovery was only slight by day 40.
    Increased plasma concentration and decreased bilary excretion of
    ouabain was also measured in male Sprague Dawley rats 10 days after a
    single oral dose (25 g/kg body weight) of TCDD, or
    1,2,3,7,8,9-hexaCDD (Yang et al., 1983b). Two other congeners,
    1,2,4,6,7,9-hexaCDD and 1,3,6,8-tetraCDD had no effect on these
    parameters.

         When hepatocytes from TCDD-treated (10 g/kg body weight) male
    Sprague Dawley rats were incubated with labelled ouabain or procaine
    amide ethobromide (PAEB) 10 days post-treatment, both the rate of
    uptake and the steady-state concentration of ouabain were decreased,
    whereas the uptake of PAEB was unaffected by TCDD (Eaton & Klaassen,
    1979). The dose of TCDD was very small relative to ouabain
    (approximately 150 mol/litre), so it is not likely that TCDD exerted
    its effect by competing with the drug for transport into bile. These
    data suggest that the hepatic membrane transport process for ouabain
    may be selectively damaged by TCDD.

         Peterson et al. (1979a) observed a positive correlation between
    the levels of hepatic plasma membrane ATPase activities, biliary
    excretion of ouabain, and bile flow in vivo after TCDD treatment.
    However, in a further experiment, using perfused rat liver, Peterson
    et al. (1979b) demonstrated that biliary excretion of ouabain and
    liver membrane ATPase activities could be decreased independently.
    Therefore, ATPase activities cannot be directly responsible for the
    reduced ouabain excretion.

    7.4.3  Porphyria

         Chronic sublethal exposure to TCDD produces an accumulation of
    porphyrins in the liver and an increase in urinary porphyrin
    excretion. In stages of manifest porphyria, accumulation of porphyrins
    occurs not only in the liver but also in the kidney and spleen
    (Goldstein et al., 1982). It was demonstrated that mice respond to
    four weekly doses of 25 g TCDD/kg body weight with hepatic porphyria,
    accompanied by an increase in aminolevulinic acid (ALA) synthetase
    activity, and liver lesions (Goldstein et al., 1973), whereas a single

    dose of 5, 25, or 100 g TCDD/kg body weight did not induce porphyria
    nor ALA synthetase activity in the rat (Woods, 1973). The suggested
    species difference was later ruled out by Cantoni et al. (1981) and
    Goldstein et al. (1976, 1982). Chronic administration of 1 g TCDD/kg
    body weight per week to rats for 16 weeks resulted in hepatic
    porphyria (Goldstein, 1976a,b, 1982). In contrast, single oral doses
    as high as 30 g TCDD/kg did not produce porphyria either acutely or
    16 weeks later. A 6-month recovery period following the final dose was
    not long enough to reverse the porphyria. Urinary porphyrins and
    hepatic ALA synthetase activity remained maximally elevated, while
    hepatic porphyrin levels decreased during this period. Failure to
    demonstrate porphyria in rats after chronic administration of TCDD for
    13 weeks (Kociba et al., 1976) or 2 years (Kociba et al., 1978) was
    suggested to be the result of unsatisfactory porphyrin analysis
    (Goldstein et al., 1982).

         To further characterize TCDD-induced porphyria, Cantoni et al.
    (1981) performed a 45-week study to follow the pattern of porphyrin
    excretion in rats exposed orally to 0.01, 0.1, or 1.0 g TCDD/kg body
    weight/week. They found an increase in the coproporphyrin level in the
    initial phase of exposure, which remained the only sign of exposure in
    the lowest-dose group. A marked porphyric state appeared only in the
    1.0 g/kg dose group, commencing 8 months after dosing started. At
    that time urinary porphyrin excretion was 70 times higher than in
    control rats. The excretion pattern was characterized by increased
    levels of carboxylated porphyrins.

         In attempts to elucidate the mechanism of TCDD-induced porphyria,
    the effects of TCDD on the enzymes involved in the synthesis and
    catabolism of porphyrins have been studied. TCDD was found to be a
    potent inducer of ALA synthetase, the initial and rate-limiting enzyme
    in haeme synthesis in the liver of chicken embryos (Poland & Glover,
    1973b). Elevated ALA synthetase activity has since been demonstrated
    also in mice and rats (Goldstein et al., 1973, 1982; Kociba et al.,
    1976, 1978). However, the TCDD-induced increase does not appear after
    acute exposure and only after several weeks of chronic exposure to
    TCDD. Jones & Sweeny (1980) failed to demonstrate increased ALA
    activity in mice exposed to 25 g TCDD/kg body weight per week for 11
    weeks, although porphyria was evident. Thus, induction of ALA
    synthetase does not seem to be the primary event in TCDD-induced
    porphyria. Elder et al. (1976, 1978) suggested that decreased hepatic
    porphyrinogen decarboxylase is the primary event in porphyria induced
    by halogenated aromatics. TCDD depresses this enzyme activity in
    vivo in the liver of mice (Jones & Sweeny, 1980; Elder & Sheppard,
    1982; Cantoni et al., 1984a,b) but not in vitro (Cantoni et al.,
    1984b).

         A decrease in porphyrinogen decarboxylase activity was present
    one week after a single dose of 75 g/kg body weight and continued to
    decrease with time, thus preceding the increase in hepatic porphyrins,
    which started to rise during the first 2 weeks after treatment (Smith
    et al., 1981). TCDD- induced depression of hepatic uroporphyrinogen
    decarboxylase activity occurs also in the newly regenerating liver, as
    demonstrated by Smith et al. (1985) in C57Bl/10 mice 10 days after
    partial (2/3) hepatectomy. The mice were treated with 75 g TCDD/kg
    body weight orally 4 weeks before hepatectomy. Greig et al. (1984)
    demonstrated that pretreatment of five different strains of mice with
    12.5 mg Fe2+ one week before the administration of 75 g TCDD/kg
    body weight had a synergistic effect on porphyria assessed 5 weeks
    after dosing, expressed as increased hepatic porphyrin and decreased
    porphyrinogen decarboxylase activity. Iron alone did not increase
    hepatic porphyrin levels, nor did it affect hepatic porphyrinogen
    decarboxylase activity.

    7.4.4  Epidermal effects

         Chloracne and associated pathological changes in the skin are
    among the most sensitive and widespread responses to TCDD in humans.
    Similar skin toxicity is expressed only in a limited number of animal
    species, namely rabbits, monkeys, and hairless mice. To characterize
    the epidermal response and to elucidate the mechanism(s) of toxicity
    to epidermal cells, numerous studies have been performed both in
    vivo and in vitro.

    7.4.4.1  In vivo studies

         The acnegenic activity of TCDD and related compounds has been
    tested in the rabbit ear bioassay, first developed by Adams et al.
    (1941) for industrial applications. The test substance is applied to
    the inner surface of one of the ears, while pure vehicle is applied to
    the other. Responses indicative of acnegenic activity include comedo
    formation, increased ear thickness, and hyperkeratosis. Mild
    irritation, increased ear thickness, slight enlargement of follicular
    aperture, slight exfoliation and slight crust formation alone are not
    considered indicative of acnegenic activity. Microscopically there is
    conversion of sebaceous cells in the hair follicles into
    keratin-forming cells. A dose-dependent, positive response was found
    in this assay when a total dose of 1, 3, or 10 g TCDD was applied on
    three successive days (Jones & Krizek, 1962). Also Schwetz et al.
    (1973) found a dose-dependent acnegenic response in the rabbit-ear
    bioassay after repeated applications of 4-40 g TCDD/ear, five days
    per week for four weeks, corresponding to total doses of 80-800 g. No
    response was obtained when the total application was 8 ng. Poiger &
    Schlatter (1980) applied a single dose of TCDD in various vehicles on
    the inner surface of the rabbit ear and followed the appearance of
    inflammation, hyperkeratosis, and chloracne. The minimum dose that
    induced skin lesions was around 1 g TCDD/ear when the vehicle was

    acetone, vaseline, or polyethylene glycol 1500 with 15% water. When
    TCDD was mixed with soil-water (2:1), or activated carbon-water (1:8)
    before application, 2-3 and 160 g TCDD/ear, respectively, were needed
    to induce lesions. Even 160 g TCDD/ear produced very small changes on
    the skin surface when applied as an activated carbon-water paste.

         Hairless mice constitute another in vivo model for studies of
    epidermal effects of TCDD. Puhvel et al. (1982) studied cutaneous
    changes induced by topical application of 0.1 g TCDD, three times per
    week for four weeks, in two strains of hairless mice (Skh:HR-1 and
    HRS/J). Epidermal hyperplasia, hyperkeratinization, loss of sebaceous
    glands and follicles, and keratin build-up in the dermal cysts were
    noted in both strains of mice. Follicular keratosis, considered the
    pathognomonic lesion in human chloracne, did not appear within 4 weeks
    of application. In the same study, follicular keratosis did develop
    after topical application of 2 mg 3,4,3',4'- tetrachlorobiphenyl, five
    times per week for 8 weeks, suggesting that follicular keratosis is an
    extension of the epidermal response and, thus, not related to
    metabolic changes in sebaceous glands. The authors considered hairless
    mice a less sensitive model for the chloracnegenic response than the
    rabbit ear bioassay. Similar findings were obtained when HRS/J mice
    were exposed to TCDD topically applied two to three times per week for
    four weeks (Knutson & Poland, 1982; Poland & Knutson, 1982). They
    found, with a total applied dose of 1.2 g TCDD, a moderate to severe
    response, including hyperplasia, hyperkeratosis of the interfollicular
    epidermis, squamous metaplasia of the sebaceous glands, and
    hyperkeratosis within the dermal cysts, but no keratosis in the
    sebaceous follicles.

         Soot or a benzene extract of soot, containing PCDDs and PCDFs
    from a PCB-containing transformer fire (Binghamton, New York, USA),
    were applied to 64.5 cm2 of the shaved, unabraded dorsal surface of
    (3 + 3) and (1 + 1) male and female New Zealand white rabbits (3.5
    kg), respectively (Silkworth et al., 1982). The dose applied was 500
    mg soot or benzene extract corresponding to 500 mg soot/kg body
    weight. Controls, one male plus one female, were exposed to activated
    carbon or benzene in corresponding amounts. Exposure lasted for 24 h
    and the observation period was 67 days. The soot produced no overt
    toxicity, no weight loss, and no histological findings in thymus,
    kidney, or skin, but hepatic centrilobular hypertrophy was found in
    both sexes. The soot extract gave rise to a reversible skin
    inflammation and hepatic centrilobular hypertrophy in the female only.
    No weight loss was recorded and the kidney, thymus, and skin were
    histologically normal at necropsy.

    7.4.4.2  In vitro studies

         Keratinocytes, the principal cell type of epidermis, form an in
    vitro model for studies of TCDD-induced hyperkeratosis both in human
    and animal-derived cell cultures. The response is analogous to
    hyperkeratinization in vivo. Newly confluent epidermal cell cultures

    exhibit proliferative properties, while the number of basal cells
    tends to decrease with increasing time of post-confluency growth.
    Thus, with the appropriate selection of culture medium and time of
    treatment, different aspects of TCDD toxicity in vivo can be
    modelled in vitro.

         A TCDD-induced response of in vitro keratinization was first
    demonstrated in XB cell cultures, an established keratinocyte cell
    line derived from a mouse teratoma, plated at high density to avoid
    spontaneous keratinization (Knutson & Poland, 1980b). Keratinization
    was dose-related and histologically similar to that which occurs
    spontaneously when XB cells are plated at low density. The epidermal
    proliferation in XB cells produced by TCDD could not be biochemically
    related to the response produced by cholera toxin, epidermal growth
    factor, or 12-O-tetradecanoylphorbol-13-acetate, other compounds
    known to affect cell proliferation in XB cells (Knutson & Poland,
    1984). Late passage XB cells, i.e. XBF cells, show increased cell
    density at saturation and a fusiform morphology at high density.
    Additionally, they have lost their ability to respond with
    keratinization upon TCDD treatment. Exposing XBF cells to TCDD
    concentrations in the range 10 to 11 X 10-8 mol/litre resulted in
    normal growth until confluency was reached by day 7. Thereafter
    TCDD-treated cultures showed a persistent decrease in cell growth and
    cell proliferation as well as changed morphology, whereas viability
    was unaffected (Gierthy & Crane, 1984). Reseeding these quiescent XBF
    cells, previously exposed to 10-9 mol/litre TCDD for 14 days,
    resulted in normal growth and proliferation until confluency. These
    TCDD-pretreated cells maintained their susceptibility to TCDD-induced
    changes in cell growth and morphology. Both XB cells (keratinization
    assay) and XBF cells (flat-cell-assay) have proved to be useful in
    vitro bioassays to measure "dioxin-like" activity of both
    environmental samples and of pure isomers (Gierthy et al., 1984;
    Gierthy & Crane, 1985a,b). Although XBF cells, a highly transformed
    variant of XB cells, seem to be less appropriate as a model for TCDD
    action on normal mammalian epithelial cell proliferation and
    differentiation, they seem to be more stable and easier to maintain
    than XB cells. Several continuous lines of human keratinocytes derived
    from neonatal foreskin (Milstone & Lavigne, 1984; Osborne et al.,
    1984) or squamous cell carcinomas (SCC) (Rice & Cline, 1984; Willey et
    al., 1984; Hudson et al., 1985, 1986) have been shown to respond to
    TCDD in nanomolar concentrations with a variety of signs that indicate
    alterations in the normal differentiation process. Stimulation of
    3H-thymidine incorporation was seen in post-confluent human
    epidermal cells derived from neonatal foreskin after exposure to TCDD
    (Milstone & Lavigne, 1984). Newly confluent human epidermal cells,
    derived from foreskin, responded to exposure for 4 days to 10 nmol
    TCDD/litre with decreased DNA synthesis, a decrease in the number of
    proliferating basal cells, decreased binding of epidermal growth
    factor (EGF), an increase in the number of differentiated cells, and
    increased envelope formation, i.e., a decrease in the proliferative

    capacity and an increase in the state of differentiation (Osborne &
    Greenlee, 1985). The decreases in small (basal) cell number and EGF
    binding were dose dependent, with EC50 values for TCDD of 2 and 1
    nmol/litre, respectively. The responses were also obtained with TCDF
    but not with 2,4-diCDD.

         The proliferation and differentiation of epidermal cells is
    normally regulated by several growth factors and hormones, e.g., EGF,
    vitamin A and hydrocortisone. Mouse hepatoma cells exposed to TCDD for
    24 h showed 20% inhibition of EGF binding (Karenlampi et al., 1983).

         Hudson et al. (1985, 1986) demonstrated that TCDD decreases, in
    a dose-dependent manner, the specific binding as well as the cellular
    uptake of EGF in cultures of human epidermal cells. The EC50 dose
    for inhibition was 1.8 nmol/litre. A similar inhibitory effect was
    obtained by TCDF, while 2,7-diCDD was inactive even at doses 100-fold
    greater. Maximal inhibition, almost 60%, of EGF binding in confluent
    SCC-12F cells exposed to 100 nmol TCDD/litre was obtained after a
    pretreatment period of 72 h. No effect was obtained when TCDD was
    added at the same time as EGF; thus TCDD did not compete for
    EGF-binding sites, neither did TCDD affect the process of
    internalizing the EGF. Further studies of the SCC-12F cell line
    revealed data suggesting that TCDD specifically reduces the high
    affinity EGF-binding sites in the basal cell population of this cell
    line (Hudson et al., 1986).

         The addition of 10-6 mol hydrocortisone/litre to the medium
    antagonized the growth inhibition of SCC cells grown in 10-10 mol
    TCDD/litre (Rice & Cline, 1984). Hydrocortisone stimulated several
    aspects of keratinocyte differentiation. These stimulatory effects
    were abolished in the presence of 10-8 mol TCDD/litre, although TCDD
    alone had no effect on these parameters. The hydrocortisone level in
    the medium was unaffected by TCDD. TCDD, and even more so
    hydrocortisone, were able to stimulate stratification in SCC cultures
    held at confluence for extended periods. This effect was opposed by
    vitamin A (Rice & Cline, 1984).

         Like TCDD, vitamin A suppressed the stimulation of keratinocyte
    differentiation by hydrocortisone. However, vitamin A had no effect on
    TCDD-induced growth inhibition or its reversal by hydrocortisone (Rice
    et al., 1983; Rice & Cline, 1984).

         Epidermal transglutaminase (ETG) activity, the marker enzyme for
    terminal differentiation, was increased by treatment of basal
    keratinocyte cultures from neonatal BALB mice with 10-9 mol
    TCDD/litre for 5 to 12 days, although morphologically no signs of
    terminal differentiation were present. A parallel increase in ETG
    activity was present when these cells were grown in medium rich in
    Ca2+, although these cells did stratify and differentiate (Puhvel et
    al., 1984).

    7.4.5  Effects on the immune system

         TCDD produces a pronounced atrophy of the thymus, spleen, and, to
    a lesser, extent the peripheral lymph nodes of most experimental
    animals. Since Buu-Hoi et al. (1972a) reported on TCDD-induced thymic
    atrophy, many studies in rats, mice, guinea-pigs, and monkeys have
    shown that the thymus is one of the organs most severely affected by
    TCDD. Lesions in the thymus appear at exposure levels well below those
    inducing lesions in other organs. Although there is species variation
    in the degree and severity of other organ effects, the effects of TCDD
    on lymphoid tissues is consistent in all species. Further
    investigations of the effect of TCDD on the immune system, which is a
    rapidly proliferating and differentiating organ system containing many
    cellular components in a highly organized and regulated network, have
    revealed that TCDD affects both the humoral-mediated (section 7.4.5.2)
    and the cell-mediated immune response (section 7.4.5.3). Also the
    complement system, a key component of the innate immunity, is affected
    by TCDD treatment (White et al., 1986). Damage to the thymus and to
    the cell-mediated immune system seems to be rather specific in that it
    occurs at doses considerably lower than those affecting other immune
    functions (Faith & Luster, 1979). Thymic involution is believed to be
    a direct effect on the gland, and not secondary to factors such as
    undernutrition (van Logten et al., 1980), altered levels of hormones
    including corticosteroids (Vos et al., 1973; van Logten et al., 1980),
    pituitary hormones (Vos et al., 1973), and thymosin (Vos et al.,
    1978a,b), or zinc deficiency (Vos et al., 1978a,b), or to a direct
    cytotoxic effect on lymphocytes (Vos et al., 1978a,b). A direct effect
    of TCDD on mouse fetal thymus organ cells grown in vitro was
    demonstrated at concentrations as low as 10-10 mol/litre (Dencker et
    al., 1985). Within the thymus, lymphocytes in the cortex, i.e., the
    immature T-cells, are more severely affected in TCDD-treated animals
    than are lymphocytes in the medulla, i.e., the mature T-cells. Thus,
    it seems that TCDD impairs the differentiation of thymocytes into
    immunocompetent T cells. Greenlee et al., (1985) obtained results
    demonstrating a direct effect of TCDD on thymus epithelial (TE) cells.

         High levels of Ah receptors (section 7.8.1) have been found in
    the thymus (Carlstedt-Duke, 1979; Mason & Okey, 1982; Gasiewicz &
    Rucci, 1984). Studies with C57Bl/6 mice (responsive to TCDD), DBA/2
    mice (less responsive to TCDD), and B6D2F1 mice, hybrid mice from
    crosses between these strains, suggest that TCDD-induced thymic
    involution, as well as immunosuppression, segregates with the Ah locus
    in these strains of mice (Poland & Glover, 1980; Clark et al., 1983;
    Vecchi et al., 1983; Nagarkatti et al., 1984; Dencker et al., 1985).
    The most profound and persistent effect of TCDD on the immune system
    is found when TCDD is administered during pre- and/or immediate
    postnatal life. Contrary to other experimental animals investigated,
    rainbow trout (Salmo gairdneri) are relatively resistant to the
    immunosuppressive effects of TCDD (Spitzbergen et al., 1986). No
    humoral-mediated effects, and only minor cell-mediated effects, were
    present at doses which caused clinical toxicity.

    7.4.5.1  Histopathology

         Lymphoid organs, primarily thymus but also spleen and lymph
    nodes, have been found to be affected by TCDD over a wide spectrum of
    dose ranges in adult rats, guinea-pigs, and mice (Gupta et al., 1973;
    Vos et al., 1973; Vos and Moore, 1974; McConnell et al., 1978b). The
    marked reduction in the size of thymus has been referred to as atrophy
    (Gupta et al., 1973; Vos & Moore, 1974), regression (Allen et al.,
    1975), or involution (Kociba et al., 1976), though these terms do not
    clearly represent the pathogenesis of this lesion but are more a
    description of the final event. Toxic effects on thymus appeared in
    adult guinea-pigs, rats, and mice exposed to eight weekly doses of 0.2
    g TCDD/kg body weight, six weekly doses of 5 g TCDD/kg body weight
    and four weekly doses of 5 g TCDD/kg body weight respectively (Vos et
    al., 1973). The thymus from moribund animals, or from animals that
    died from TCDD exposure, showed a dose-dependent decrease in the
    number of cortical lymphocytes, markedly smaller thymic lobules, and
    loss of demarcation between the cortex and medulla. Guinea-pigs, the
    species most severely affected by TCDD, showed large cystic Hassall
    bodies, filled with polymorphonuclear leukocytes (Gupta et al., 1973;
    Vos et al., 1973). Guinea-pigs that received lethal doses of TCDD
    showed scattered necrosis of lymphocytes in the cortical region with
    concomitant phagocytosis by macrophages as early as 5 days
    post-exposure (McConnell et al., 1978b). The effect was more apparent
    at day 14, and by day 20 it was difficult to differentiate the cortex
    from the medulla. At day 20 little evidence of necrosis remained,
    though karyorrhectic debris and prominent phagocytosis indicated that
    this had occurred. Since the thymus from guinea-pigs surviving the
    TCDD dose for 30 days was usually normal microscopically (Vos et al.,
    1973; McConnell et al., 1978b), it seems that thymic necrosis must be
    an early event in the course of the toxic syndrome. Furthermore, in
    animals which survive, thymic regeneration seemed to be rapid.
    Decreased weight of thymus, loss of cortical cells, and cell necrosis
    have also been found in hamsters exposed to TCDD (Gasiewicz et al.,
    1986).

         Depletion of lymphoid cells in the spleen, intestinal tract, and
    various lymph nodes observed in guinea-pigs, rats, and mice (Gupta et
    al., 1973; McConnell et al., 1978b) was less extensive than in the
    thymus. The major effect in the spleen of rats is the loss of the
    T-cell-dependent areas, namely the periarterial lymphoid sheet and the
    paracortical areas (Vos & Moore, 1974).

         Depressed immunoglobulin levels were reported for 1- and
    4-month-old C57Bl/6 mice exposed to four and six weekly doses,
    respectively, of 25 g TCDD/kg body weight (Vos & Moore, 1974).
    Feeding 10, 20, 50, or 100 g TCDD/kg in the diet depressed
    dose-dependently the Y-globulin level in 7-week-old Swiss-Webster mice
    (Hinsdill et al., 1980).

         Vos & Moore (1974) demonstrated a dose-related lymphocyte
    depletion of thymus cortex, spleen, and intestinal lymph nodes in
    maternally exposed pups of rats and mice. However, no effect on
    immunoglobulin levels was observed in 25-day-old rats maternally
    exposed (5 g TCDD/kg body weight) on days 0, 7, and 14. The
    developing lymphoid tissues were found to be more sensitive to TCDD
    than were the lymphoid tissues of adults or young.

    7.4.5.2  Humoral-mediated immunity

         The humoral-mediated immunity (Tables 51 and 52) operates through
    antibody-producing cells and is transferable by serum. This system
    includes classical antibody-mediated protective immunity and immediate
    hypersensitivity reactions. Vos et al. (1974) reported a significant
    decrease in the alpha-, -, and gamma-globulin levels in C57BL/6 mice
    given non-toxic doses of TCDD. The effects of TCDD on specific humoral
    immunity responses in adult animals are summarized in Table 51.
    Feeding levels of 10 g TCDD/kg body weight or more reduced the
    primary and secondary antibody response to both sheep red blood cells
    (sRBC) and tetanus toxin in male Swiss-Webster mice (Hinsdill et al.,
    1980). Weekly oral doses of 1 or 10 g TCDD/kg body weight reduced
    both the primary and secondary serum antitetanus titres in male New
    Zealand rabbits (Sharma et al., 1984). The secondary, but not the
    primary, serum tetanus antitoxin level was decreased in Hartley
    guinea-pigs given eight weekly doses of 0.2 g TCDD/kg body weight
    (Vos et al., 1973). Results presented by Vecchi et al. (1980, 1983)
    show that single doses as low as 1.2 g TCDD/kg body weight to C57BL/6
    mice decreased the number of plaque-forming spleen cells in response
    to an injection of the thymus-dependent antigen sRBC. The response was
    dose dependent and lasted for at least 42 days. Luster et al. (1985)
    found a decrease in anti-sRBC plaque-forming spleen cell production in
    B6C3F1 and DBA/2 mice 5 days after single oral doses of 2 and 10 g
    TCDD/kg body weight, respectively. A dose of 30 g TCDD/kg body weight
    was needed to produce a significant antibody response to the
    thymus-independent antigen type III pneumococcal polysaccharide (sIII)
    in C57Bl/6 mice (Vecchi et al., 1980). With sRBC and the
    thymus-independent antigen trinitrophenylated Brucella abortus,
    Clark et al. (1981) found a depressed number of spleen plaque-forming
    cells in C57BL/6 mice only with a total dose of 40 g/kg body weight
    given as four equal weekly doses. Chastain & Pazdernik (1985)
    demonstrated decreased numbers of plaque-forming spleen and bone
    marrow cells in response to the thymus-independent antigen
    trinitrophenylated lipopolysaccharide (LPS). Spleen and bone marrow
    cells were collected from male C57BL/6 mice 7 days after a single ip
    dose of 30, 60, 90, or 120 g TCDD/kg body weight. The decrease
    occurred at lower doses in the spleen cell assay but was more
    pronounced at higher doses in the bone marrow cell assay. The number
    of antibody-producing cells, measured asplaque-forming cells,
    following immunization with either sRBC or LPS was reduced in B6C3F1
    mice treated with 1 and 5 g TCDD/kg body weight (Tucker et al.,
    1986).



    
    Table 51. Effects of TCDD on humoral-mediated immune responses in adult animals
                                                                                                                             

    Species/strain      Sex/age/weightf     TCDD exposure                 Parameter measurede
    (reference)                             Frequency/route/dose
                                                                                                                             

    Mice
       Swiss-Webster    F/4-7 weeks/NR      fed 10, 20, 50, 100, or       primary and secondary sRBCa antibody
                                            500 g/kg continuously        level,
    (Hinsdill et al.,                       in the diet for 5 weeks       primary and secondary serum tetanus
      1980)                                                               antitoxin level

       C57BL/6J         M/6-8 weeks/NR      single ip dose of             anti-sRBCa plaque-forming spleen cells
                                            1.2, 6, or 30 g/kg body      anti-SIIIb plaque-forming spleen cells
    (Vecchi et al.,                         weight
      1980)

       C57BL/6J         M/6-8 weeks/NR      four weekly ip doses of       anti-sRBCa plaque-forming spleen cells
                                            0.1, 1, or 10 g/kg body      anti-TNP-BAc plaque-forming spleen cells
    (Clark et al.,                          weight
      1981)

       C57BL/6          M/8-10 weeks/NR     single ip dose of             anti-sRBCa plaque-forming spleen cells
       C3H/HeN                              1.2, 6, or 30 g/kg body
       DBA/2                                weight
       AKR
       B6D2F1

    (Vecchi et al.,
      1983)

       C57BL/6          M/6-8 weeks/NR      single ip dose of 30, 60,     anti-TNP-LPSd plaque-forming spleen cells
                                            90, or 120 g/kg body weight  anti-TNP-LPSd plaque-forming bone marrow
    (Chastain &                                                           cells
      Pazdernik, 1985)
                                                                                                                             

    Table 51. (contd - 2)
                                                                                                                             

    Species/strain      Sex/age/weightf     TCDD exposure                 Parameter measurede
    (reference)                             Frequency/route/dose
                                                                                                                             
    Mice (contd).
       DBA/2N           F/6-8 weeks/        single oral doses of 5,       anti-sRBCa plaque-forming spleen cells
                        18-21 g             10, or 50 g/kg body weight
    (Luster et al.,
      1985)

       B6C3F1           F/6-8 weeks/        single oral doses of 0.2,     anti-sRBCa plaque-forming spleen cells
                        18-21 g             1, 2, 5, or 10 g/kg body
    (Luster et al.,                         weight
      1985)

    Guinea-pigs
       Hartley          F/NR/256 g          eight weekly oral doses of    primary serum tetanus antitoxin level
                                            0.008, 0.04, 0.2, or          secondary serum tetanus antitoxin level
    (Vos et al., 1973)                      1.0 g/kg body weight

    Rabbit
       New Zealand      M/Adult/3 kg        eight weekly oral doses of    primary and secondary serum tetanus
                                            0, 0.01, 0.1, 1.0, or 10.0    antitoxin level
    (Sharma et al.,                         g/kg body weight
      1984)
                                                                                                                             

    a     sRBC = sheep red blood cell.
    b     SIII = type III pneumoccal polysaccharide.
    c     TNP-BA = trinitrophenylated Brucella abortus.
    d     TNP-LPS = trinitrophenylated lipolysaccharide.
    e     = all parameters measured were decreased except for primary serum tetanus antitoxin level in guinea-pigs
         (Vosal., 1973).
    f     M = male;  F = female;  NR = not reported.

    Table 52. Effects of TCDD on humoral-mediated immune responses in maternally exposed animals

    Species/strain        Time of TCDD exposure         Route/dose                  Parameter measured-response
    (reference)
                                                                                                                            
    Rats
       Fisher-344 N       Prenatal day 18 and/or        oral/5 g/kg body weight    primary and secondary BGGa
                          postnatal days 0, 7, and 14                               antibody level - no effect

    (Faith & Moore,
       1977)

       Fisher/Wistar      Prenatal day 18 and/or        oral/5 g/kg body weight    primary and secondary BGGa
                          postnatal days 0, 7, and 14                               antibody level - no effect

    (Faith & Luster,
       1979)

    Mice
       Swiss-Webster      four weeks before mating,     in the diet/                primary and secondary sRBCb
                          throughout gestation and      1, 2.5, 5, 10, or 20 /kg   antibody level - no effect;

    (Thomas &             lactation                                                 primary anti sRBCb plaque
       Hinsdill, 1979)                                                              forming spleen cells - decreased
                                                                                                                            


    a     = BGG - bovine gamma-globuline.
    b     = sRBC - sheep red blood cell.
    


         Only minor effects on antibody responses have been reported in
    rodents maternally exposed to TCDD (Table 52). Single oral doses of 5
    g TCDD to pregnant Fisher-344 N (poor immunological responder) and
    Fisher-Wistar rats (good immunological responder) on gestation day 18
    and/or postnatal days 0, 7, and 14 did not affect the antibody reponse
    to bovine gammaglobulin (BGG) in the offspring (Faith & Moore 1977;
    Faith & Luster, 1979).

         Dietary exposure of female Swiss-Webster mice to 2.5 or 5 g
    TCDD/kg for 4 weeks before mating and throughout gestation and
    lactation resulted in normal antibody production in the offspring but
    in a decrease in anti-sRBC plaque-forming spleen cells.

         In vitro exposure of B6C3F1 spleen cells to 10-9 mol
    TCDD/litre decreased the production of anti-sRBC plaqueforming cells
    (Luster et al., 1984). The ED50 for this effect was found to be 7
    nmol/litre when TCDD was present from the first day of culture (Tucker
    et al., 1986). Similar activity to that of TCDD was found with
    2,3,7-triCDD and 1,2,3,7,8-pentaCDD, whereas 2,8-diCDD and octaCDD
    were without effect at concentrations up to 5 x 10-8 mol/litre
    (Tucker et al., 1986).

         The doses producing 50% suppression of splenic IgM response to
    sRBC in C57Bl/6 mice were 7.1 and 85 g/kg body weight, respectively,
    for 1,2,3,6,7,8-hexaCDD and 1,2,3,4,6,7,8-heptaCDD given as single
    oral doses 2 days prior to sRBC challenge (Kerkvliet et al., 1985).
    OctaCDD had no effect even at doses of 100 and 500 g/kg body weight.

         Humoral immune responses of the rainbow trout (Salmo
    gairdneri) were not significantly impaired even at doses of TCDD
    that caused clinical toxicity (Spitsbergen et al., 1986).

    7.4.5.3  Cell-mediated immunity

         Cell-mediated immunity (CMI) operates through specifically
    sensitized lymphocytes and is transferred by these cells. Processes
    included in this system are classical cell-mediated protective
    immunity (which protects against fungi, bacteria, and viruses),
    delayed type hypersensitivity, rejection of tumors and foreign tissues
    such as transplants, and graft versus host response. Many assays, both
    in vivo and in vitro, have been developed to test CMI functions.
    Besides a reduction in the number of immunologically competent cells
    after TCDD exposure (Gupta et al., 1973; Vos et al., 1973; Zinkl et
    al., 1973) TCDD has been demonstrated to induce a decreased CMI
    response in adults (Table 53) and even more in maternally exposed
    animals (Table 54). Delayed hypersensitivity response, correlatingwith
    decreased host resistance to infectious agents in man, was depressed
    in rodents exposed to low levels of TCDD (Vos et al., 1973; Faith &
    Moore, 1977; Sharma et al., 1978; Faith & Luster, 1979; Thomas &

    Hinsdill, 1979; Hinsdill et al., 1980; Clark et al., 1981).
    TCDD-exposure also adversely affects host susceptibility to bacteria,
    viruses, tumor cells and endotoxins (Thigpen et al., 1975; Thomas &
    Hinsdill, 1979; Hinsdill et al., 1980; Luster et al., 1980; Clark et
    al., 1983).

         Depressed graft versus host response was reported in 2-month-old
    C57BL/6 mice exposed to 4 weekly oral doses of 5 g TCDD/kg body
    weight (Vos et al., 1973), whereas no effect was seen in a subsequent
    study on 1- and 4-months old C57BL/6Sch mice (Vos & Moore, 1974). In
    the same study, Fisher-344 rats maternally exposed to TCDD showed
    decreased graft versus host response and prolonged allograft-rejection
    time. The latter effect was also demonstrated in maternally
    TCDD-exposed C57BL/6Sch mice (Vos et al., 1973). Proliferative
    responses of spleen and/or thymus lymphocytes, stimulated by mitogens
    specific for the generation of B-lymphocytes and/or T-lymphocytes from
    TCDD-exposed animals, were depressed both in adults (Vos & Moore,
    1974; Sharma et al., 1978) and maternally exposed rodents (Vos &
    Moore, 1974; Faith & Moore, 1977; Vos et al., 1978; Faith & Luster,
    1979; Luster et al., 1980). However, Thomas & Hinsdill (1979) found no
    effect on the lymphoproliferative response in offspring from
    Swiss-Webster mice fed with up to 5 g TCDD/kg diet 4 weeks before
    mating and throughout gestation and lactation. Depressed
    lymphoproliferative response is regarded as an extremly sensitive
    indicator of immunotoxicity, rather than as a predictor of immune
    dysfunction. Cytotoxic T-cell generation in response to allogeneic
    antigens has been demonstrated in male DBA/2, C57Bl/6, and B6D2F1 mice
    given four weekly ip injections of 1 ng/kg body weight (Clark et al.,
    1981, 1983). At this dose no effects were seen on delayed
    hypersensitivity, antibody response, thymus cellularity, or enzyme
    induction. The adverse effect of TCDD on CMI function seems to be an
    age-related phenomenon in rodents. In order to obtain a complete and
    persistent immune suppression, TCDD exposure must occur during
    ontogenesis of the immune system. In the initial experiments on the
    developing immune system, Vos & Moore (1974) exposed Fisher-344 rats
    to 1 mg TCDD/kg body weight on gestation days 11 and 18 and on
    postnatal days 4, 11, and 18, or to 5 g TCDD/kg body weight on
    postnatal days 0, 7, and 14. CMI functions adversely affected included
    in vitro immune competence of spleen and thymus lymphoid cells,
    delayed hypersensitivity reaction, prolonged allograft-rejection times
    and reduced graft versus host activity. The immune suppression
    demonstrated persisted throughout the study, i.e., for 145 days. The
    depression of T-cell-dependent immune functions appeared to occur
    without helper-cell function being affected (Faith & Moore 1977).

         Attempts to study direct effects of TCDD on lymphocytes in
    vitro were previously hampered by the low solubility of TCDD in
    physiological buffers (Matsumura & Benezet, 1973). Vos & Moore
    (1974)obtained no lymphoproliferative response in unstimulated or PHA-
    or concanavalin-A-stimulated rat thymus organ cells and mouse spleen
    cells when cultured in the presence of up to 20 ng TCDD/ml. Dencker et

    al. (1985) demonstrated that mouse fetal thymus cells cultured in
    vitro in the presence of TCDD gave a similar response similarly to
    that occurring in vivo, i.e., with a dose-dependent inhibition of
    the time-dependent increase in the number of lymphoid cells
    (EC50=10-10 mol TCDD/litre). It could not be determined with
    certainty whether the decreased cell number caused by TCDD was due to
    reduced cell proliferation or to increased cell death. The
    TCDD-induced suppression of mitogen-stimulated lymphoproliferation has
    recently been demonstrated to be mediated by thymus epithelial cells
    (Greenlee et al., 1985).

    7.4.5.4 Macrophage function

         The primary pathway of endotoxin detoxification is thought to be
    macrophage-dependent. Thus the increased sensitivity to endotoxin
    following TCDD treatment (Vos et al., 1978; Thomas & Hinsdill, 1979)
    was suggestive of macrophage dysfunction. However, the number of
    peritoneal macrophages, as well as their capacity to mediate cytolytic
    and cytostatic effects, was not adversely affected by single ip doses
    of 1.2, 6 or 30 g TCDD/kg body weight to male C57Bl/6J mice
    (Mantovani et al., 1980), nor was the ability of macrophages to reduce
    nitroblue tetrazolium affected by four to five weekly oral doses of 50
    g TCDD/kg body weight in Swiss-Webster mice (Vos et al., 1978a).
    Macrophage function does not appear to be altered by TCDD.

    7.4.6  Myelotoxicity

         TCDD treatment inhibits the bone marrow haematopoiesis in mice,
    both in vivo and in vitro, by directly altering colony growth of
    stem cells (Luster et al., 1980, 1985; Chastain & Pazdernik, 1985).
    The bone marrow granulocyte-macrophage progenitor cell (CFU-GM)
    production was reduced in B6C3F1 mice, receiving 1 g TCDD/kg body
    weight, but was unaffected in DBA/2 mice, even at a dose of 50 g
    TCDD/kg body weight (Luster et al., 1985). Chastain & Pazdernik (1985)
    used the B-lymphocyte colony-forming unit assay to demonstrate
    reductions in spleen and bone marrow B-cell function of C57BL/6 mice
    exposed to single ip doses of TCDD in the range 30 to 120 g TCDD/kg
    body weight. Bone marrow B-cells tended to be more suseptible to TCDD
    than spleen B-cells. A maternal oral dose of 5 g TCDD/kg body weight
    on the 14th day of gestation and postnatal days 1, 7, and 14 resulted
    in a weak normocytic anaemia indicative of depressed erythrogenesis,
    decreased bone marrow cellularity, and decreased stem cell
    proliferation in B6C3F1 offspring, assessed 7 days after weaning
    (Luster et al., 1980). In vitro exposure of B6C3F1 bone marrow cells
    to 10-9 mol TCDD/litre resulted in decreased CFU-GM development and
    decreased number of erythrocyte colony-forming units. The decrease in
    CFU-GM occurred 1 day post-treatment and remained below control values
    until 10 days post-treatment (Luster et al., 1985).



    
    Table 53. Effects of TCDD on cell-mediated immunity responses in adult animals
                                                                                                                            

    Species/strain      Sex/age/weightd        TCDD exposure                 Parameter measured-responsed
    (reference)                                Frequency/route/dose
                                                                                                                            

    Rats
       CD               F/NR/185 g             six weekly oral doses of      delayed hypersensitivity to tuberculin - NE
                                               0.2, 1.0, or 5.0 g/kg
    (Vos et al., 1973)                         body weight

    Mice
       C57BL/6          M/6-8 weeks/NR         four weekly ip doses of       delayed hypersensitivity to sRBCa - Dec;
                                               0.1, 1.0, or 10.0 g/kg       delayed hypersensitivity to oxazalone - Dec;
    (Clark et al.,                             body weight                   generation of alloantigen-specific cytotoxic
       1981)                                                                 T-cells - Dec

       C57BL/6          M/NR/NR                four weekly ip doses of       resistance to Herpes virus challenge - Dec;
                                               0.001, 0.01, 1.0, or          generation of alloantigen-specific cytotoxic
    (Clark et al.,                             10.0 g/kg body weight        T-cells - Dec
       1983)

       DBA/2            M/NR/NR                four weekly ip doses of       generation of alloantigen-specific cytotoxic
                                               0.001, 0.01, 0.1, 1.0, or     T-cells - Dec
    (Clark et al.,                             10.0 g/kg body weight
       1983)

       Swiss-Webster    F/4-7 weeks/NR         five weeks feeding of diets   resistance to Salmonella typhimurium challenge - Dec;
                                               containing 10, 50, or         resistance to Listeria monocytogenes challenge - Dec;
    (Hinsdill et al.,                          100 g/kg body weight         contact sensitivity to 2,4-dinitro-1-fluoro-
       1980)                                                                 benzene - Dec
                                                                                                                              

    Table 53 (contd - 2)
                                                                                                                            

    Species/strain      Sex/age/weightd        TCDD exposure                 Parameter measured-responsed
    (reference)                                Frequency/route/dose
                                                                                                                            
    Mice (continued)
       C57BL/6J         M/6-8 weeks/NR         single ip doses of 1.2,       number of peritoneal macrophages - Dec;
                                               6 or 30 g/kg body weight     number of splenic natural killer cells - Dec;
    (Mantovani                                                               macrophage mediated cytolysis - NE;
       et al., 1980)                                                         macrophage mediated cytostasis - NE

       C57BL/6          M/NR/NR                four weekly ip doses of       generation of allospecific cytotoxic T-cells - Dec
       DBA                                     0.001 g/kg body weight       - NE
       B6D2F1                                                                - Dec

    (Nagarkatti
       et al., 1984)

       CD-1             M/NR/28                two, four, or eight weekly    lymphoproliferative response of PHAb- and PWMc-
                                               doses of 0.01, 0.1, 1.0, or   stimulated splenic cells - Dec
    (Sharma &                                  10.0 g/kg body weight
       Gehring, 1979)

       C57BL/6Jfh       M/4 weeks/NR           four weekly oral doses of     resistance to Salmonella bern challenge - Dec;
                                               0.5, 1.0, 5.0, 10.0, or       resistance to Herpes virus challenge - NE
    (Thigpen et al.,                           20.0 g/kg body weight
       1975)

       C57BL/6          M/2 months/            four weekly oral doses of     graft versus host activity - Dec
                        24.4 g                 0.2, 1.0, 5.0, or 25.0
    (Vos et al.,                               g/kg body weight
       1973)

       C57BL/6Sch       M/1 month/NR           four weekly oral doses of     lymphoproliferative response of PHAb-stimulated
                                               1.0, 5.0, or 25.0 g/kg       spleen cells - Dec;
    (Vos & Moore,                              body weight                   graft versus host activity - NE
       1974)
                                                                                                                            

    Table 53 (contd - 3)
                                                                                                                            

    Species/strain      Sex/age/weightd        TCDD exposure                 Parameter measured-responsed
    (reference)                                Frequency/route/dose
                                                                                                                            
    Mice (continued)
       C57BL/6Sch       M/4 months/NR          six weekly oral doses of      lymphoproliferative response PHA-stimulated
                                               1.0, 5.0, or 25.0 g/kg       spleen cells - NE
    (Vos & Moore,                              body weight                   graft versus host activity - NE
       1974

       Swiss            M/3-4 weeks/NR         four or five weekly oral      resistance of Listeria monocytogenes - NE;
                                               doses of 50 g/kg body        number of peritoneal macrophages - NE;
    (Vos et al.,                               weight                        macrophage reduction of nitroblue tetrazodium - NE
       1978a)

    Guinea-pigs
       Hartley          F/NR/256 g             eight weekly oral doses       delayed hypersensitivity to tuberculin - Dec
                                               of 0.008, 0.04, 0.2, or
    (Vos et al, 1973)                          1.0 g/kg body weight

    Rabbits
       New Zealand      M/adult/3 kg           eight weekly oral doses       delayed hypersensitivity to tuberculin - Dec
                                               of 0.01, 0.1, 1.0, or
    (Sharma et al.,                            10.0 g/kg body weight
       1984)
                                                                                                                            

    a     sRBC = sheep red blood cells.
    b     PHA = phytohaemagglutinin.
    c     PWM = poke weed.
    d     NR = not reported;  NE = no effect;  Dec = decreased;  M = male;  F = female.

    Table 54.  Effects of TCDD on cell-mediated immunity responses in maternally exposed animals
                                                                                                                            

    Species/strain      TCDD exposure                                  Age when     Parameter measured - responsef
    (reference)         Frequency/route/dose                           testedg
                                                                                                                            

    Rats
       Fisher/Wistar    5 g/kg body weight on gestation day 18        25 days      lymphoproliferative response of
                        and on postnatal days, 0, 7, and 14, or                     PHAa- and ConAb-stimulated spleen
    (Faith & Luster,    5 g/kg body weight on postnatal days                       and thymus cells - Dec;
      1979)             0, 7, and 14                                                delayed hypersensitivity to
                                                                                    tuberculin - Dec

       Fisher-344       5 g/kg body weight on gestation day 18        25 days      lymphoproliferative response of
                        and on postnatal days 0, 7, and 14, or                      PHAa- and Conb-stimulated spleen
    (Faith & Moore,     5 g/kg body weight postnatal days 0, 7,                    and thymus cells - Dec;
      1977)             and 14                                                      delayed hypersensitivity to
                                                                                    oxazolone - Dec

       Fisher 344       1 g/kg body weight on gestation days 11       25 days      lymphoproliferative response of
                        and 18 and on postnatal days 4, 11 and 18 or                PHAa-stimulated spleen cells and of
    (Vos & Moore,       5 g/kg body weight on postnatal days 0,                    PHAa- and ConAb-stimulated thymus-
      1974)             7 and 14                                                    cells - Dec

    Mice
       B6C3F1 c         1, 5, or 15 g/kg body weight on               NR           lymphoproliferative response of
                        gestation day 14 and on postnatal days 1,                   mitogen-stimulated spleen cells:
    (Luster et al.,     7, and 14                                                   PHAa - Dec;
      1980)                                                                         ConAb - Dec;
                                                                                    LPSd - NE;
                                                                                    Macrophage proliferation - NE;
                                                                                    Phagocytizing ability - NE;
                                                                                    Resistance to Listeria monocyto-
                                                                                    genes - Dec;
                                                                                    Resistance to PYB6 tumor cells - Dec
                                                                                                                            

    Table 54 (contd).
                                                                                                                            

    Species/strain      TCDD exposure                                  Age when     Parameter measured - responsef
    (reference)         Frequency/route/dose                           testedg
                                                                                                                            
    Mice (continued)
       Swiss-Webster    Feeding 1, 2.5, or 5 g TCDD/kg                5-6 weeks    Contact sensitivity to 2,4-dinitro-
                        4 weeks before mating, throughout                           1-fluorobenzene - Dec;
    (Thomas &           gestation, and 3 weeks postnatally                          lymphoproliferative response of
      Hinsdill, 1979)                                                               PHAa- and ConAb-stimulated spleen
                                                                                    and thymus cells - NE;
                                                                                    Resistance to Salmonella typhimurium
                                                                                    endotoxin - Dec; Resistance
                                                                                    to Listeria monocytogenes - NE

       C57BL/6Sch       2 or 5 g/kg body weight on gestation days     23 days      Skin graft assay - prolonged skin
                        14 and 17 and on postnatal days 1, 8, and 15                graft rejection time
    (Vos et al., 1974)

       Swiss            10 g/kg body weight on postnatal days         22 days      lymphoproliferative response of
                        1, 4, 8, 11, 15, and 18                                     PHAa-, ConAb- &
    (Vos et al.,                                                                    PWMe-stimulated thymus cells - Dec
      1978a)
                                                                                                                            

    a     PHA = phytohaemagglutinin.
    b     ConA = concanavalin A.
    c     B6C3F1 = progeny to female C57BL/6N and male C3H mice.
    d     LPS = lipopolysaccharide.
    e     PWM = poke weed.
    f     Dec = decreased,  NE = no effect.
    g     NR = not recorded.
    


    7.4.7  Effects on the intermediary metabolism

         Changes in intermediary metabolism have been demonstrated in
    TCDD-treated experimental animals. The circulating concentration of
    glucose in TCDD-treated rats was decreased relative to ad
    libitum-fed (Zinkl et al., 1973; Schiller et al., 1985) and pair-fed
    (Gasiewicz et al., 1980; Potter et al., 1983) control rats.
    TCDD-treated and pair-fed controls (both groups were schedule fed) had
    similar concentrations of serum glucose (Christian et al., 1986b). The
    TCDD-induced hypoglycaemia was not caused by altered pancreatic
    function, as judged by insulin and glucagon levels (Potter et al.,
    1983).

         Reduced hepatic glycogen content, compared to the value in
    control rats, was reported in Sprague Dawley rats 16 days after a
    single ip dose of 20 g TCDD/kg body weight (Weber et al., 1983).
    However, in studies by Christian et al. (1986b), compared to pair-and
    schedule-fed control rats, TCDD-treated (75 g/kg body weight daily)
    Sprague Dawley rats had significantly increased hepatic glycogen
    levels but unaffected cardiac and muscle (gastrocnemius) levels of
    glycogen 2-8 days post-treatment.

         TCDD-treated rats maintained a normal overall nitrogen balance,
    as judged by urinary urea, creatinine, and ammonia levels, but
    exhibited changes in certain plasma protein levels (Christian et al.,
    1986b).

         Elevated circulating cholesterol levels were found in
    TCDD-treated rats (Albro et al., 1978; Poli et al., 1980; Schiller et
    al., 1985), whereas circulating free fatty acids and triacylglycerols
    were decreased in TCDD-treated rats, when compared to pair-and
    schedule-fed control rats (Christian et al., 1986b).

         A marked accumulation of hepatic lipid has been found in rats
    after single doses of TCDD (Cunningham & Williams, 1972; Gupta et al.,
    1973; Albro et al., 1978; Schiller et al., 1985). At a sublethal dose
    of TCDD, hepatic triglyceride and free fatty acid levels were elevated
    already one day after dosing. Abnormal lipid deposition patterns
    persisted for at least 2 months (Albro et al., 1978). The increased
    level of triglycerides in the liver (Schiller et al., 1985; Christian
    et al., 1986b) of TCDD-treated rats was not accompanied by increases
    in cardiac or muscle (gastrocnemius) triacylglycerol levels (Christian
    et al., 1986b). Hepatic lipid synthesis in Wistar rats, measured as
    the 1-h incorporation of 3H-acetate, was not affected by TCDD
    treatment when studied 7 days after exposure to 10 g TCDD/kg body
    weight (Cunningham & Williams, 1972).

         Mice (both C57BL/6 and DBA/2 strains) responded to TCDD treatment
    with dose-dependent decreases in serum levels of glucose, cholesterol,
    and triglycerides, and increases in hepatic triglycerides, whereas
    serum glycerol and free fatty acid levels were unaffected (Chapman &
    Schiller, 1985).

         Hartley guinea-pigs given a lethal ip dose of 2 g TCDD/kg body
    weight had significantly increased circulating levels of cholesterol
    esters, triglycerides, and phospholipids but a normal free fatty acid
    level 7 days post-treatment. The increase in serum lipids was
    accompanied by a pronounced increase in low density lipoproteins,
    particularly the very low density lipoprotein fraction (Swift et al.,
    1981). The plasma cholesterol and triglyceride levels were elevated
    also when TCDD-treated guinea-pigs were compared to pair-fed controls
    (Gasiewicz & Neal, 1979).

         A TCDD-induced increase in plasma triglyceride levels was also
    noted in New Zealand rabbits, fed 20 g TCDD/kg body weight, whereas
    plasma cholesterol was unaffected 12 weeks after dosing (Lovati et
    al., 1984). TCDD treatment did not alter the liver lipid levels (free
    and esterified cholesterol, triglycerides, and phospholipids), but the
    triglyceride level was significantly increased. These results were
    valid both for rabbits fed normal chow and those fed a
    cholesterol-rich (0.5%) diet. Golden Syrian hamsters exposed to 1000
    g TCDD/kg body weight, orally or ip, had elevated plasma cholesterol
    levels until 20 days after exposure but normal levels on day 50,
    whereas serum triglyceride levels were normal until day 20 and then
    became significantly lower than those of controls (Olson et al.,
    1980b).

    7.4.8  Enzyme induction

         Primarily, TCDD has been found to increase enzyme activities
    although observations on enzyme inhibition have also been made. Since
    the first reports of enzyme systems as targets for TCDD (Buu-Hoi et
    al., 1971a, 1972b; Greig, 1972; Poland & Glover, 1973b,c), enzyme
    induction has become the most extensively studied biochemical response
    produced by TCDD. The mixed function oxidase system (MFO), capable of
    metabolizing both endogenous and foreign lipophilic compounds to more
    polar products, has been the most thoroughly investigated one, and
    arylhydrocarbon hydroxylase (AHH) and 7-ethoxy-resorufin 0-deethylase
    (EROD) are the most frequently assayed enzyme activities in this
    system. TCDD has also been reported to affect UDP-glucuronosyl
    transferases (UDPGT) (Thunberg et al., 1980, 1984) and
    glutathione-S-transferases (GT) (Manis & Apap, 1979), which are
    multifunctional enzyme systems involved in conjugating a wide variety
    of compounds.

         Most studies have been performed with microsomal enzymes, but
    TCDD has also been found to have effects on enzymes in the cytosolic
    fraction. It seems that TCDD produces organ-specific effects, and
    although, quantitatively, hepatic enzyme induction is of more concern
    than extrahepatic enzyme effects, the latter may qualitatively be as
    important. Studies in different species have revealed that enzyme
    induction due to TCDD exposure also is a species-specific phenomenon.

         Time course studies have shown that maximal increases in enzyme
    activities are reached within 3 to 4 days post-treatment. After a lag
    period of about 2 to 3 weeks, enzyme activities begin to return to
    normal levels (Hook et al., 1975a,b; Lee & Suzuki, 1980; Lucier et
    al., 1973; Poland & Glover, 1973a).

         According to Kitchin & Woods (1979), TCDD-induced AHH activity
    did not reach the normal level until 6 months after rats were exposed
    to a single daily dose of 2 g/kg body weight.

         Hook et al. (1975a) found no apparent dependence on age when
    studying AHH induction in CD rats which were 10 to 335 days old at the
    time of exposure to 25 g TCDD/kg body weight. 

         Several investigators have studied the relative potency of various 
    PCDDs and PCDFs to induce AHH and/or EROD activities (Bradlaw et al., 
    1980; Poland et al., 1976; Bandiera et al., 1984a,b; Sawyer & Safe, 
    1985; Mason et al., 1986).  They found an apparent structure-activity 
    relationship between the location of the halogen atoms on the 
    dibenzo-p-dioxin molecule and the ability to induce AHH activity 
    both in vivo and in vitro.  Isomers with halogens at the four 
    lateral ring positions produced a greater biological response than 
    those with halogens at three lateral ring positions, while two 
    lateral halogen atoms seemed to be insufficient to produce a 
    biological response. TCDD was the most potent enzyme inducer of the 
    compounds tested.

         On a molecular basis TCDD is the most potent MFO-inducing
    compound known and MFO induction seems to be the most sensitive
    biochemical response produced by this chemical. According to Kitchin
    & Woods (1979), induction in the rat takes place after a single daily
    dose of only 0.002 g TCDD/kg body weight. In the guinea-pig (the
    animal most sensitive to TCDD toxicity), MFO induction has been
    observed, but the induced activities were low even at lethal doses
    (Hook et al., 1975a). Neither is there a correlation in cell cultures
    between induction of MFO and toxicity. Furthermore, it is known that
    metabolites of TCDD are less toxic and more readily excreted than the
    parent compound (see section 8.1.5). Thus, TCDD-induced MFO activities
    represent a detoxification process rather than one leading to toxic
    effects.

         However, induction of MFO activities might potentiate the
    toxicity of other foreign compounds requiring metabolic transformation
    by the MFO system before they can exert their toxic effect. A number
    of studies have shown that induction of MFO activities alters the

    metabolism of the model xenobiotic, benzo(a)pyrene by increasing the
    rate of microsomal metabolism, changing the metabolic profile to more
    toxic metabolites, and increasing the extent of covalent binding to
    liver microsomes (Berry et al., 1976, 1977; Uotila et al., 1978).
    TCDD, applied topically or subcutaneously increased the
    carcinogenicity of 3-methyl cholanthrene (MC) in DBA/2 mice (Kuori,
    1978), but decreased the carcinogenicity of
    7,12-dimethylbenz(a)anthracene (DMBA) in CD-1 mice (DiGiovanni et al.,
    1979a). These authors suggested that TCDD induces the MFO system and
    thus increases activation of MC to the ultimate carcinogen, as well as
    inactivation of DMBA, which would explain these effects.

         Furthermore, increased MFO activities might adversely affect
    important metabolic pathways of endogenous compounds. The effects of
    TCDD on enzyme activities, both MFO and others, involved in such
    biological pathways as keratinization, steroid metabolism, lipid
    metabolism, plasma membrane function and porphyrin metabolism, are
    discussed under separate sections. The minute quantities of TCDD
    required for maximal enzyme induction or suppression, the long
    duration of the effect, and the stereospecific requirements suggest a
    specific interaction of TCDD with a cellular species, possibly at the
    gene level. Accordingly considerable research has been directed toward
    the study of the genetic regulation of AHH induction by TCDD. A
    hepatic cytosolic species that bound TCDD has been suggested as the
    receptor for the hepatic AHH activity. Numerous studies of this
    cytosolic receptor in several species and tissues have been performed
    and it seems that there is a structural gene, the Ah locus, for this
    receptor, which is responsible for the expression of various enzyme
    activities (see sections 7.8 and 7.8.1).

    7.4.8.1 Studies on rats

         The effect of TCDD on enzyme activities has been most extensively
    investigated in the rat. In the liver, TCDD has been shown to increase
    both the content of cytochrome P-450 (Lucier et al., 1973, 1986;
    Poland & Glover, 1974a,b; Hook et al., 1975a; Aitio & Parkki, 1978;
    Kitchin & Woods, 1979; Madhukar & Matsumura, 1981; Goldstein & Linko,
    1984) and cytochrome b5 (Lucier et al., 1973; Hook et al., 1975a), as
    well as the microsomal enzyme activities involved in the oxidative
    transformation and conjugation of xenobiotics, e.g., aniline
    hydroxylase, arylhydrocarbon hydroxylase (AHH), biphenyl hydroxylase,
    7-ethoxycoumarin-0-deethylase (ECOD), EROD, and UDPGT. These enzyme
    activities have been investigated in a vast number of studies, some of
    them quoted in Table 55. Goldstein & Linko (1984) demonstrated that
    TCDD induced two isozymes of cytochrome P-450 (P-448) in the liver but
    only one of these in extrahepatic tissues of young Sprague Dawley rats
    2 days after a single oral dose of 25 g/kg body weight.

        Table 55. Studies demonstrating in vivo induction of mixed function
    oxidases and UDP-glucuronosyltransferases in TCDD-exposed strains of rats

                                                                                  
                              Rat
    Enzyme activity           strain          Reference
                                                                                  
    Aniline hydroxylase       SD       Beatty et al. (1978)

                              CD       Lucier et al. (1973); Hook et al. (1975a)

    Aryl hydrocarbon          SD       Poland & Glover (1973a, 1974a); Beatty et
    hydroxylase                        al. (1978); Manis & Apap (1979);
                                       Haaparanta et al. (1983); Thunberg et al.
                                       (1984); Lucier et al. (1986); Ahlborg et
                                       al. (1987)

                              CD       Lucier et al. (1973); Hook et al.
                                       (1975a); Kitchin & Woods (1979)

                              Wistar   Nagayama et al. (1983); Keys et al.
                                       (1985); Bannister et al. (1986); Farrell
                                       & Safe (1986); Mason et al. (1986);
                                       Tsyrlov et al. (1986)

    Biphenyl hydroxylase      CD       Hook et al. (1975a,b,); Kitchin & Woods
                                        (1979)

    7-Ethoxycoumarin-O-       Wistar   Aitio & Parkki (1978)
    deethylase

    7-Ethoxyresurofin-O-      SD       Haaparanta et al. (1983)
    deethylase

                              CD       Kitchin & Woods (1979)

                              Wistar   Keys et al. (1985); Bannister et al.
                                       (1986); Farrell & Safe (1986); Mason et
                                       al. (1986)

    UDP-glucuronosyl-
    transferase:
       p-nitrophenol          SD       Thunberg et al. (1980, 1984); Ahlborg et
                                       al. (1987)

                              CD       Lucier et al. (1973, 1975a, 1986); Hook
                                       et al. (1975a)

    Table 55 (Cont.)  Studies demonstrating in vivo induction of mixed
    function oxidases and UDP-glucuronosyltransferases in TCDD-exposed
    strains of rats
                                                                                    
                            Rat
    Enzyme activity         strain                   Reference
                                                                                    
    UDP-glucuronosyl-
    transferase:
       p-nitrophenol          Wistar   Aitio et al. (1979); Thunberg & Hkansson
          (cont'd.)                    (1983)

       o-aminophenol          Wistar   Aitio et al. (1979)

       4-methylumbelli-
       ferone                 Wistar   Aitio & Parkki (1978)
                                                                                    
    
         Microsomal glutathion-s-transferase (GT) did not respond to
    TCDD (Aitio & Parkki, 1978; Mukitari et al., 1981; Baars et al.,
    1982), but cytosolic GT was induced both by a single dose of 17 g
    TCDD/kg body weight 2 days post-treatment (Manis & Apap, 1979), and by
    near lethal or lethal doses 1 and 6 days after dosing (Mukitari et
    al., 1981; Baars et al., 1982; Hassan et al., 1983). Glutathione
    reductase was also increased, while glutathione peroxidase, both total
    and Se-dependent, and the content of reduced glutathione were reduced
    by TCDD treatment (Hassan et al., 1983, 1985a,b,c).

         The following hepatic enzyme activities involved in drug
    metabolism have been reported to be unaffected by TCDD treatment in
    the rat: N-and O-demethylation (Lucier et al., 1973; Poland &
    Glover, 1973a; Hook et al., 1975a; Beatty et al., 1978; Kitchin &
    Woods, 1979; Madhukar & Matsumura, 1981), epoxide hydratase (EH)
    (Aitio & Parkki, 1978), -glucuronidase (Lucier et al., 1973, 1975)
    and NADPH cytochrome c reductase (Poland & Glover, 1974; Aitio &
    Parkki, 1978; Kitchin & Woods, 1979; Madhukar & Matsumura, 1981). The
    glucuronide conjugation of bilirubin (Aitio et al., 1979), estrone,
    and testosterone (Lucier et al., 1975a) by liver microsomes from
    TCDD-treated rats was not different when compared to control rats.

         Some hepatic enzyme activities not belonging to the MFO system,
    which are affected by TCDD treatment include aldehyde dehydrogenase
    (Deitrich et al., 1978; Lindahl et al., 1978), delta-aminolevulinic
    acid synthetase (see section 7.4.3), DT-diaphorase (Beatty & Neal,
    1977), transglutaminase (see section 7.4.4), ornithine decarboxylase
    (Nebert et al., 1980; Potter et al., 1982; Farrell & Safe, 1986),
    plasma membrane ATPases (see section 7.4.2.2), porphyrinogen

    carboxylase (see section 8.4.3), prostaglandin synthetase (see section
    7.4.9), enzymes involved in testosterone metabolism (see section
    7.4.9), and RNA polymerase (Kurl et al., 1982).

         Prenatal and postnatal exposure via milk to TCDD, at doses of 3
    g/kg body weight to pregnant rats on days 5, 10 and 16 of gestation,
    induced hepatic AHH and UDPGT activities in the offspring. The effect
    was seen 8 days post-partum and persisted for at least 2 weeks. The
    inductive effect was due both to exposure to TCDD via milk and to the
    activation of an inducing mechanism after birth. Fetal liver AHH was
    slightly increased during late gestation, although the UDPGT activity
    and the cytochrome P-450 content were not (Lucier et al., 1975b).

         Administration of 2.5 g TCDD/kg body weight to pregnant rats on
    day 17 of gestation increased the AHH- and N-hydroxylation activities
    and cytochrome P-450 content in the fetal liver on day 20 of gestation
    (Berry et al., 1976).

         AHH induction due to TCDD has been reported to occur also in the
    brain (Hook et al., 1975a), kidney (Poland & Glover, 1973a; Hook et
    al., 1975a; Aitio & Parkki, 1978; Potter et al., 1982; Nagayama et
    al., 1983), lung (Poland & Glover, 1973a; Hook et al., 1975a; Aitio &
    Parkki, 1978; Nagayama et al., 1983), prostate (Lee & Suzuki, 1980;
    Haaparanta et al., 1983; Nagayama et al., 1983), thymus (Nagayama et
    al., 1983), and intestine (Poland & Glover, 1973a; Hook et al.,
    1975a), but intestinal AHH activity was found to be unaffected by 17
    and 20 g TCDD/kg body weight (Aitio & Parkki, 1978; Manis & Apap,
    1979). Testicular (Poland & Glover, 1973a; Hook et al., 1975a; Aitio
    & Parkki, 1978) and adrenal (Guenthner et al., 1979) AHH activities
    were not induced by sublethal doses of TCDD. The O-deethylation
    activity in kidney, lung, and prostate was increased, but no effect
    was seen on the activity in testes or intestine (Aitio & Parkki, 1978;
    Haaparanta et al., 1983). UDPGT activities in kidney, lung, intestine,
    and brain were increased, while no effect was seen on testicular UDPGT
    (Hook et al., 1975a). Similar results were reported by Aitio & Parkki
    (1978), though in their study intestinal UDPGT was not affected.

         Renal biphenyl hydroxylation activity has been found to increase
    after TCDD treatment, but no effect on this enzyme activity was seen
    in lung, intestine, brain, or testes (Hook et al., 1975a). Elevated
    levels of cytochrome P-450 were found in prostate (Lee & Suzuki, 1980)
    and mammary gland (Rikans et al., 1979), but not in adrenals
    (Guenthner et al., 1979). Less testicular cytochrome P-450 was found
    after a single dose of 25 g TCDD/kg body weight (Tofilon & Piper,
    1982). The GSH tranferase activity was increased in the lung but not
    in kidney, intestine, testes (Aitio & Parkki, 1978), or prostate (Lee
    & Suzuki, 1980).

         Neither EH nor NADPH cytochrome c reductase were inducible by
    TCDD in kidney, lung, intestine, testes (Aitio & Parkki, 1978),
    mammary gland (Rikans et al, 1979), or prostate (Lee & Suzuki, 1980).
    The ED50 values for hepatic AHH and EROD induction were determined
    in immature male Wistar rats 13 days after a single ip dose of
    2,3,7-triCDD, TCDD, 1,3,7,8-tetraCDD, 1,2,3,7,8-pentaCDD,
    1,2,4,7,8-pentaCDD or 1,2,3,4,7,8-hexaCDD (Mason et al., 1986). The
    order of enzyme-inducing capacity was TCDD > 1,2,3,7,8-pentaCDD >
    1,2,3,4,7,8-hexaCDD > 1,2,4,7,8-pentaCDD > 2,3,7-triCDD >
    1,3,7,8-tetraCDD (Table 56).

    7.4.8.2  Studies on mice

         Enzyme induction studies in mice have been performed mainly with
    two strains genetically separated at the Ah locus, thus making them
    responsive (C57Bl/6 (B6)) or non-responsive (DBA/2 (D2)) to induction
    of hepatic cytochrome P-450-related enzyme activities by aromatic
    hydrocarbons, e.g., 3-methyl-cholanthrene (3-MC). However, the
    extraordinary potency of TCDD for enzyme induction revealed increased
    hepatic cytochrome P-450 content as well as AHH and O-deethylase
    activities both in B6 and D2 mice after sublethal exposure to TCDD
    (Poland & Glover, 1974a,b,c; Jones & Sweeney, 1977; Greenlee & Poland,
    1978). Studies of MFO induction in five responsive and five
    non-responsive strains of mice by Poland & Glover (1974a,b,c) revealed
    that there were no consistent differences between the strains when
    considering the extent to which TCDD induced AHH, O-deethylase,
    N-demethylase and O-demethylase activities in the liver, kidney,
    lung, skin, or bowel. The ED50 for hepatic AHH induction was
    determined to be 10-9 mol/kg body weight in the responsive strain
    and > 10-8 mol/kg body weight in the non-responsive strain (Poland &
    Glover, 1975). Fully induced hepatic AHH activity was obtained both in
    responsive (C57Bl/6 and AKR/Qdj) and non-responsive (DBA/2 and DDD)
    strains of mice 3 days after an ip dose of 30 g TCDD/kg body weight
    (Nagayama et al., 1985a). Both AHH and EROD were induced in C57Bl/6
    mice 7 days after an ip dose of 0.32 g/kg body weight (Bannister et
    al., 1986). Both hepatic AHH and ornithine decarboxylase activities
    were similarly induced in C57Bl/6 and DBA/2 mice after a single ip
    dose of 100 g TCDD/kg body weight, but at 2 g TCDD/kg body weight
    these enzymes were induced only in the C57Bl/6 strain (Nebert et al.,
    1980). Two daily doses of 0.1 g TCDD, topically applied, increased
    the epidermal AHH activity in two strains of hairless mice (Puhvel et
    al., 1982). Contrary to observations in rats, TCDD induces testicular
    AHH activity both in B6 and D2 mice 40h after an ip dose of 50 g/kg
    body weight (Mattison & Thorgeirsson, 1978).



    
    Table 56. Structure activity relationships for some PCDDs
                                                                                                                             

    PCDD                        In vitro EC50 values (M)a            In vivo ED50 values (mmol/kg)bLD50 c
    Congener                                                                                              

                          Receptor       AHH            EROD          AHH       EROD       Body      Thymic       Guinea-pig
                          binding                                                          weight    atrophy      g/kg
                                                                                                                  body
                                                                                                                  weight
                                                                                                                             

    1-                  > 1.0x10-4  > 1.0x10-4   > 1.0x10-4
    2,8-                  3.2x10-6  > 1.0x10-4   > 1.0x10-4                                                       > 300 000
    1,2,4-                1.3x10-5    4.8x10-5     2.2x10-6
    2,3,6-                2.2x10-7
    2,3,7-                7.1x10-8    3.6x10-7     1.4x10-7           19.6      19.6                  98.1           29 400
    1,2,3,4-              1.3x10-6    3.7x10-6     2.4x10-6
    1,2,3,8-                          6.1x10-7
    1,3,7,8-              7.9x10-7    5.9x10-7     3.2x10-7           31.2      77.6       132       100
    2,3,6,7-              1.6x10-7    6.1x10-8     1.1x10-8
    2,3,7,8-              1.0x10-8    7.2x10-11    1.9x10-10          0.004     0.003      0.05      0.09                 2
    1,2,3,4,7-            6.4x10-6    6.6x10-7     8.2x10-7
    1,2,3,7,8-            7.9x10-8    1.1x10-8     1.7x10-8           0.031     0.056      0.62      0.17                 3.1
    1,2,4,7,8-            1.1x10-6    2.1x10-8     1.1x10-8           2.82      0.56       34        11.2              1125
    1,2,3,4,7,8-          2.8x10-7    2.1x10-9     4.1x10-9           0.03      0.130      1.63      1.07                72.5
    1,2,3,6,7,8-c         5.7x10-10   3.1x10-8                                                                             70-100
    1,2,3,7,8,9-c         1.4x10-9    4.6x10-8                                                                             60-100
    1,2,3,4,6,7,8c                    1.3x10-7                                                                           > 600
    1,2,3,4,6,7,9-c                   3.7x10-6
    1,2,3,4,6,7,8,9-    > 1.0x10-5  > 1.0x10-4   > 1.0x10-4
                                                                                                                             

    a     Estimated concentrations needed to displace 50% of 3H-TCDD bound to liver cytosol receptor from Wistar
          rats and to produce 50% maximum enzyme induction in the rat hepatoma 11-4-II E cell line (Bradlaw & Casterline,
          1979; Mason et al., 1986).
    b     Studies in immature male Wistar rats (Mason et al., 1986).
    c     McConnell et al., 1978b.
    


    7.4.8.3  Studies on guinea-pigs

         The guinea-pig, the species most sensitive to the toxic effects
    of TCDD, does not respond with liver toxicity or with  extensive
    enzyme induction.

         Hook et al. (1975a) investigated the effect of a single oral dose
    of 0.175 g TCDD/kg body weight on Hartley guinea-pig MFO and UDPGT
    activities in liver, kidney, and lung. AHH induction was found only in
    the kidney. In none of the tissues was there an effect on UDPGT
    activity. The biphenyl 4-hydroxylase activity was increased in all
    tissues, whereas hepatic biphenyl 2-hydroxylase was decreased. With
    three daily doses of 1 g TCDD/kg body weight, Hassan et al. (1983)
    found an increase in hepatic AHH 6 days post-treatment. They also
    found slightly increased in vitro lipid peroxidation but no effect
    on glutathione content or on the enzyme activities facilitating
    peroxidation, reduction, or transfer of glutathione. The DT-diaphorase
    activity was not affected by a single oral dose of 0.6, 3.0, or 6.0 mg
    TCDD/kg body weight (Beatty & Neal, 1977). The maximal increases were
    4.4 and 22 times for AHH and EROD activities, respectively. The
    testicular cytochrome P-450 content in Hartley guinea-pigs was
    decreased by 50% one day after a single dose of 1 mg TCDD/kg body
    weight. The effect persisted for at least 9 days (Tofilon, 1980). No
    effect was seen on microsomal haeme content or on the activities of
    NADPH-cytochrome C reductase and sorbitol dehydrogenase, the marker
    enzyme for testicular protein synthesis. Thus, the decrease in
    cytochrome P-450 induced by TCDD does not seem to be a nonspecific
    inhibition of protein synthesis.

    7.4.8.4  Studies on rabbits

         Studies on enzymes in rabbits have been performed in the New
    Zealand albino strain exposed to single doses of 10 to 30 mg TCDD/kg
    body weight for 1 to 5 days. Both in adults (Johnson &
    Muller-Eberhard, 1977a,b) and in neonates exposed in utero (Norman
    et al., 1978; Kohli & Goldstein, 1981), TCDD increased the content of
    cytochrome P-450. It also induced the formation of immunologically
    distinct cytochromes P-450 in adult and neonatal liver (Norman et al.,
    1978). Increased cytochrome P-450 was observed in the kidney but not
    in the lung (Liem et al., 1980; Kohli & Goldstein, 1981). Renal and
    pulmonary cytochrome P-450 reductase, investigated by Liem et al.
    (1980), were not affected by TCDD treatment. Data on MFO induction and
    suppression are conflicting. Liem et al. (1980) reported increased AHH
    and O-deethylase activities in lung and kidney, whereas Hook et al.
    (1975a) saw no effect on the AHH activity in the lung and reported a
    decrease in hepatic AHH activity after a single oral dose of 0.5 g
    TCDD/kg body weight. Biphenyl 4-hydroxylase induction was seen in the
    liver by Johnson et al. (1979). Hook et al. (1975a) detected such
    induction in lung, but no effect in the liver and kidney (Hook et al.

    1975a). Furthermore, Hook et al. (1975a) reported no effects on
    biphenyl-2-hydroxylation and UDPGT activities in liver, kidney, or
    lung. A decrease in hepatic, but not in renal or pulmonary
    N-demethylation, was found by these authors.

    7.4.8.5 Studies on hamsters

         Golden Syrian hamsters are among the animals most resistant to
    acute lethal effects induced by TCDD. Although the liver is a target
    tissue, hepatic enzyme induction has barely been studied in this
    species. When given an oral dose of 200 g/TCDD/kg body weight for a
    period of 3 days, increased hepatic glutathione-S-transferase and
    glutathione reductase activities were found, but no effects were seen
    on AHH or glutathione peroxidase activities. Neither the hepatic level
    of glutathione nor the in vitro lipid peroxidation were affected
    (Hassan et al., 1983). The ED50 values for induction of hepatic ECOD
    and reduced NAD(P), menadione oxidoreductase activities, and
    cytochrome P-450 content in male Golden Syrian hamsters were 1.0, 2.0,
    and 0.5 g TCDD/kg body weight, i.e. extremely low doses as compared
    to doses that produce tissue damage and lethality in this species
    (Gasiewicz et al., 1986).

    7.4.8.6 Studies on cows

         Three dairy Holstein cows (500-600 kg) received a single oral
    dose of 0.05 (two cows) or 7.5 g TCDD/kg body weight (Jones et al.,
    1986). The cow receiving the high dose was killed on day 7 and those
    receiving the low dose on day 14. AHH and EROD activities were
    markedly induced in the high-dose but not in the low-dose animals.

    7.4.8.7  Studies on chick embryos

         AHH and delta-aminolevulinic acid synthetase in the chick embryo
    have been reported to be extremely sensitive to the inductive effects
    of TCDD (Poland & Glover, 1973b,c). Maximal induction occurred with
    155 pmol TCDD/egg. The induction was relatively long lasting, with 70%
    of the maximum induced activity present 5 days following a single dose
    of TCDD. Structure-activity studies demonstrated a good correspondence
    between the toxicity and induction potency of a series of
    dibenzo-p-dioxin congeners (Poland & Glover, 1973c).

         ED50 values for the induction of hepatic microsomal EROD, AHH,
    and 4-dimethylaminoantipyrine-N-demethylase in 2-week-old white
    Leghorn cockerels on day 5 after TCDD exposure were 778, 302, and 561
    ng/kg body weight, respectively, and aldrin epoxidase was inhibited by
    TCDD treatment (Sawyer et al., 1986). Hepatic and cardiac EROD
    activities were increased in white Leghorn chicken embryos exposed to
    TCDD in ovo at doses between 1000 and 10 000 pmol/egg (Quilley &
    Rifkind, 1986).

    7.4.8.8 Studies on cell cultures

         TCDD has a very low toxicity in cell cultures, yet it is a very
    potent inducer of AHH activity in these systems, including lymphocytes
    and primary hepatocytes, as well as established and transformed cell
    lines.

         The inducibility of lymphocyte AHH has been investigated in
    mitogen-stimulated human lymphocytes from the venous blood of healthy
    volunteers. Kouri et al. (1974) found a dose-dependent increase in AHH
    activity (0, 0.1, 1.0, 10, or 100 ng TCDD/ml medium for 24 h). The
    optimal dose was about 10 ng/ml, and the maximal induction was by a
    factor of 2 to 3. On the contrary, Gurtoo et al. (1979) found no
    dose-response correlation, in the dose range 1.7 to 20 ng TCDD/ml,
    when measuring lymphocyte AHH induction. To circumvent the limitation
    of prior mitogen activation when studying AHH induction in
    lymphocytes, Freedman et al. (1979) used the human B-lymphocyte
    RPMI-1788 cell line, which does not require prior activation for the
    induction of AHH activity. The optimal concentration to stimulate AHH
    activity was determined to be 10 ng/ml medium. Highly variable
    induction of AHH (between 3- and 28- fold) was obtained by Nagayama et
    al. (1985b) in human lympho-blastoid cell lines derived from the
    peripheral blood of healthy volunteers of both sexes and of variable
    ages. The cells were exposed to 7.5 ng TCDD/ml medium for 48 h.

         In a study by Niwa et al. (1975), the estimated ED50 values for
    AHH induction by TCDD in 11 established cell lines, in fetal primary
    cultures from five animal species and cultured human lymphocytes,
    ranged from 0.04 ng/ml medium in C57Bl/6 mouse cultures and 0.08 ng/ml
    in the rat hepatoma H-4-IIE cell line to more than 66 ng/ml in the HTC
    rat hepatoma cell line. TCDD was demonstrated to be the most potent
    AHH inducer out of 24 chlorinated dibenzo-p-dioxin analogues
    (Bradlaw et al., 1980) tested in a rat hepatoma cell culture extremely
    sensitive to AHH induction, the ED50 being about 0.5 pg/106 cells.
    A 165-fold increase in AHH-activity and a 54-fold increase in EROD
    activity were obtained in rat hepatoma H-4-IIE cells when exposed to
    2 x 10-10 mol TCDD/litre for 3 days (Keys et al., 1986). In this
    system co-exposure of  TCDD with 1,3,6,8-tetraCDF and 2,4,6,8-tetraCDF
    reduced the TCDD-induced enzyme induction, whereas co-exposure of TCDD
    and TCDF resulted in an additive effect on enzyme induction (Keys et
    al., 1986). The EC50 values for AHH and EROD induction in the same
    cell system varied over 7 orders of magnitude for 14 different PCDDs
    (Table 56), the most potent being TCDD and the least potent being
    2,3,6-triCDD (Mason et al., 1986). A 2-to 650-fold AHH induction was
    observable in 8 of 22 different cell cultures exposed to 10-9 mol
    TCDD/litre for 24 h (Knutson & Poland, 1980a). The cells were derived
    from tissues and/or species susceptible to TCDD toxicity in vivo.
    Nanomolar concentrations of TCDD induced AHH activity in keratinocyte
    cultures of human (Willey et al., 1984) and animal origin (Knutson &
    Poland, 1980a).

         Five human squamous carcinoma cell lines derived from tumours of
    the epidermis and tongue responded to TCDD with increased
    O-deethylase activity, the EC50 being 10-10 to 10-9 mol/litre
    (Hudson et al., 1983a).

         Steward & Byard (1981) treated primary hepatocytes isolated from
    Sprague Dawley rats, for 48 h with various concentrations of TCDD.
    They found a 2-fold induction of the AHH activity with 3 pg TCDD/106
    cells. Maximal induction occurred with 2.4 ng TCDD/106 cells. Primary
    hepatocytes, isolated from adult male Wistar rats, exhibited a linear
    increase (from 2-to 4-fold) in AHH activity when exposed to TCDD in
    the range 10-11 to 10-8 mol/litre for 72 h (Jansing & Shain,
    1985). Primary hepatocytes from TCDD-treated (5, 10, or 25 g TCDD/kg
    body weight) rats, isolated 2 to 30 days post-treatment, showed
    decreased ouabain and alpha-aminoisobutyric acid uptake as well as
    tyrosine aminotransferase activity (Yang et al., 1983a). Treatment of
    rats with 25 mg 1,3,6,8-tetraCDD/kg body weight did not affect these
    parameters. Neither could these effects be demonstrated in primary
    hepatocytes from control rats that were treated with TCDD (50, 100, or
    200 nmol/litre medium) in vitro for 48 h.

         The induction of AHH and EROD activities of a complex PCDD/PCDF
    mixture from a fly ash extract has been reported (Safe et al., 1987).

    7.4.9 Endocrine effects

         Human exposure to TCDD has resulted in hirsutism and chloracne
    (Table 64), symptoms that suggest an alteration in endocrine
    regulation. Furthermore, chronic exposure to TCDD impaired
    reproduction in experimental animals, possibly by interfering with the
    estrous cycle (Kociba et al., 1976; Allen et al., 1977; Barsotti et
    al., 1979; Murray et al., 1979). The ability of TCDD to mimic natural
    steroids with steroid-like actions has prompted studies on the binding
    of TCDD to steroid hormone receptors.

         Over-production of glucocorticoids mimics some of the symptoms of
    TCDD toxicity, e.g., involution of lymphoid tissues, oedema, and
    mobilization of fatty acids from adipose tissues. Thus TCDD might
    increase glucocorticoid activity by binding to glucocorticoid
    receptors. However, TCDD was unable to displace 3H-dexamethasone, a
    potent synthetic glucocorticoid, from the normal rat cytosol
    glucocorticoid receptor even when present in 200-fold molar excess
    (Neal et al., 1979). Poland et al. (1976) demonstrated that cortisol
    and synthetic glucocorticoids did not bind to the TCDD receptor. An
    increase in the plasma level of corticosterone was found in male
    Sprague Dawley rats 7 and 14 days after a single oral dose of 50 g
    TCDD/kg body weight (Neal et al., 1979). With the method used, a
    variety of fluorescent adrenocortical steroid hormone derivatives was
    measured. In contrast, Balk & Piper (1984), using a competitive
    binding radioassay for corticosterone, reported decreased blood levels

    (29 and 26% of controls) of corticosterone in male Sprague Dawley rats
    on days 14 and 21, respectively, after a single oral dose of 25 g
    TCDD/kg body weight. Accumulation of 11--hydroxy-progesterone in the
    blood of TCDD-treated rats was noticed on day 14 (Balk & Piper, 1984).
    Neal et al. (1979) reported 100% mortality within 6 days in
    adrenalectomized rats given 10, 20, 40, or 80 g TCDD/kg body weight.
    Adrenalectomy and hypophysectomy could not prevent liver lesions,
    reduced growth rate, or thymic involution in female Fisher-344 rats
    given a single oral dose of 10 or 20 g TCDD/kg body weight (van
    Logten et al., 1980). Thymic effects of TCDD became even more severe
    after hypophysectomy. Daily sc injections of 0.25 mg growth hormone
    had a positive influence on body weight gain but did not protect
    against thymic involution in hypophysectomized rats.

         Single oral or interperitoneal doses of TCDD between 7 and 100
    g/kg body weight decreased the serum thyroxine (T4) level in the rat
    (Bastomsky, 1977; Potter et al., 1983, 1986b; McKinney et al., 1985;
    Pazdernik & Rozman, 1985; Rozman et al., 1985b), but not in the
    guinea-pig, after a single oral dose of 2 g/kg body weight (McKinney
    et al., 1985). The serum triiodothyronine (T3) level in TCDD-treated
    rats was reported to be increased (Bastomsky, 1977; Potter et al.,
    1986b), unaffected (Potter et al., 1983), or decreased (Pazdernik &
    Rozman, 1985; Rozman et al., 1985b). Increased serum thyrotropin (TSH)
    (Bastomsky, 1977; Potter et al., 1986b), increased thyroid
    131I-uptake and increased biliary excretion of T4, but not of T3
    (Bastomsky, 1977) have been reported in TCDD-treated rats.

         Rats treated with TCDD at doses up to about 30 mg TCDD/kg body
    weight and pair-fed controls had similar serum T4, T3, and TSH levels
    when compared to ad libitum fed control rats (Potter et al., 1983,
    1986b). Thus, it is unlikely that hypophagia is responsible for the
    TCDD-induced changes in serum thyroid hormone levels. When mature
    TCDD-treated rats were compared to pair-fed controls, there were no
    functional alterations in thyroid status or thermogenesis, including
    increased serum levels of T3, T4, and TSH after acute cold
    challenge, increased total oxygen consumption after moderate cold
    exposure or decreased basal metabolic rate as compared to ad
    libitum fed control rats (Potter et al., 1986b). A significant
    hypothermia was, however, observed in young TCDD-exposed rats
    receiving 45 g/kg body weight as a single ip dose (Potter et al.,
    1983).

         Available data on serum T4, T3, and TSH levels are not
    sufficient to state whether TCDD-treated rats are functionally
    hypothyroid, euthyroid, or hyperthyroid.

         In ovo exposure of white Leghorn chicken embryos to TCDD in
    the dose range 1 to 10 000 pmol/egg increased the cardiac release of
    prostaglandins (Quilley & Rifkind, 1986). Potter et al. (1983)
    reported decreased levels of insulin in serum and pancreas and of

    somatostatin in the gastric antrum of Sprague Dawley rats 7 days after
    a single ip dose of 45 g TCDD/kg body weight, when compared to
    pair-fed control rats. The somatostatin levels in serum, liver, and
    pancreas were not affected, neither was the serum glucagon level.

         The finding that steroids are an endogenous substrate for the
    hepatic MFO system (Kuntzman et al., 1965) suggests that compounds,
    such as TCDD, that influence the activity of this enzyme system may
    alter steroid metabolism in vivo, and consequently also the
    magnitude of steroid-mediated functions. As demonstrated by Gustafsson
    & Ingelman-Sundberg (1979), the metabolic profiles of 4-androstene-3,
    17-dione, 5 alpha-androstane-3 alpha, 17 -diol and 4-pregnene-3,
    20-dione in hepatic microsomes from SD rats, treated with 20 g
    TCDD/kg body weight for 4 consecutive days, were changed when compared
    to control rats 1 day post-treatment. The changes were most pronounced
    in female rats. When five daily doses of 1 g TCDD/kg body weight were
    given to pregnant rats for 12 or 13 days during gestation, hepatic
    microsomes showed decreased ability to form catechol estrogens and to
    hydroxylate testosterone. However, this decrease did not relate to
    altered circulating estradiol levels (Shiverick & Muther, 1983). TCDD
    treatment did not affect the glucuronidation of testosterone and
    estrogen (Lucier et al., 1975a) or prostaglandin synthesis  (Kohli &
    Goldstein, 1981). TCDD decreased the hepatic and uterine estrogen
    receptor levels in 25-day-old Long-Evans rats 2 to 10 days after
    treatment with single ip doses of 20 or 80 g/kg body weight (Romkes
    et al., 1987). Decreased estrogen receptor levels in the liver and
    uterus also occurred 2 days after treatment with 1,2,7,8-tetraCDD,
    1,2,3,7,8- pentaCDD, and 1,2,4,7,8-pentaCDD, but dose-response
    relationships were present only for TCDD and 1,2,3,7,8-pentaCDD. The
    estradiol-induced increases in hepatic and uterine estrogen receptor
    levels were counteracted by simultaneous TCDD treatment, although the
    effect of TCDD was not dose dependent. Hepatic hydroxylation of
    testosterone in 2 -and 16 alpha- positions was not affected in male
    Sprague Dawley rats (190-200 g) 3 days prior to a single oral dose of
    15 g TCDD/kg body weight (Hook et al., 1975b). Somewhat younger male
    Wistar rats (100 g) treated with a single dose of 0.06 mmol TCDD/kg
    body weight exhibited increased levels of 7 alpha-hydroxytestosterone
    and decreased levels of 3alpha-, 16alpha- and 16 -hydroxylated
    testosterone, as well as of androstenedione, in hepatic microsomes
    when compared to control rats (Keys et al., 1985). Decreased
    cytochrome P-450 content was found in guinea-pig (Tofilon, 1980) and
    rat (Tofilon & Piper, 1982) testes for at least one week after oral
    TCDD treatment of 1 g/kg and 25 g/kg body weight, respectively.
    Testicular AHH activity was induced by TCDD in mice (Mattison &
    Thorgeirsson, 1978), but not in rats after a single oral dose of 20 g
    TCDD/kg body weight (Aitio & Parkki, 1978). Mittler et al. (1984)
    studied the effect of single ip doses of 0.2, 1, or 5 g TCDD/kg body
    weight on testicular 16-alpha-testosterone hydroxylase (16-TH),
    6--hydroxytestosterone (6-HT), and 7-alpha-hydroxy testosterone
    (7-HT) activities in young Sprague Dawley rats 90 h after exposure.

    Seminiferus 16-TH activity was increased from a non-detectable level
    to 0.07-0.14 pg/mg protein and interstitial 6-HT activity was
    increased 4-fold in TCDD-treated animals. The 7-HT activity was not
    affected by TCDD, neither in the seminiferous tubules nor in the
    interstitial fraction. Serum testosterone and dihydrotestosterone were
    depressed dose-dependently by TCDD treatment, the ED50s being about
    15 g/kg body weight in male Sprague Dawley rats, when compared to
    pair-fed and ad libitum fed controls (Moore et al., 1985). The
    plasma clearance and biliary excretion of 3H-testosterone was not
    affected in Sprague Dawley rats treated with 100 g TCDD/kg body
    weight before an i.v. injection of 3H-testosterone, neither did
    castrated rats with implanted testosterone-leaking capsules respond to
    a dose of 15 or 100 g TCDD/kg body weight with decreased accessory
    sex organ weights (Moore & Peterson, 1985). Increased AHH and
    O-deethylase activities and cytochrome P-450 content have been
    reported in rat prostate after single ip and oral doses, respectively,
    of 10 g/TCDD/kg body weight (Haaparanta et al., 1983; Lee & Suzuki,
    1980).

    7.4.10  Vitamin A storage

         Decreased hepatic vitamin A storage has been reported in animals
    exposed to various chlorinated aromatic compounds (Table 57). Compared
    to other chlorinated hydrocarbons for which this effect has been
    evaluated, TCDD is more potent in its ability to reduce the vitamin A
    content of the liver.

         A single oral dose of 10 g TCDD/kg body weight decreased both
    the total amount and the concentration of vitamin A in the liver of
    adult male Sprague Dawley rats (Thunberg et al., 1979). The decrease
    was evident 4 days after dosing and progressed with time. After 8
    weeks the treated animals had a total liver vitamin A content
    corresponding to 33% of that of controls. Decreased dietary intake of
    vitamin A could not account for this difference. In a four-week study
    TCDD was given as a single oral dose of 0, 0.1, 1.0, or 10 g per kg
    body weight to adult male Sprague Dawley rats fed ad libitum with
    pelleted diets containing 1.2 (low), 3.0 (normal), or 6.0 (high) mg
    vitamin A/kg diet (Thunberg et al, 1980). Both the concentration and
    the total amount of vitamin A were decreased in a dose-dependent
    manner in the animals receiving the high vitamin A diet. In the
    animals on the normal and low vitamin A diets, significant differences
    were seen only at doses of 1.0 and 10 g TCDD/kg body weight. A
    significant increase in the UDPGT activity was observed in all dietary
    groups treated with 1.0 and 10 g TCDD per kg body weight, suggestive
    of an increased excretion of vitamin A conjugated with glucuronic acid
    (Thunberg & Hakansson, 1983). However, no correlation between the
    UDPGT activity and the reduction of hepatic vitamin A levels was seen
    when homozygous Gunn rats lacking inducible UDPGT were treated with a
    single oral dose of 20 g/kg body weight (Aitio et al., 1979) nor in
    heterozygous Gunn rats with inducible UDPGT after a single oral dose
    of 10 g/kg body weight (Thunberg & Hakansson, 1983).



    
    Table 57. The potency of various chlorinated cyclic hydrocarbons to reduce hepatic vitamin A content in the rat
                                                                                                                            
    Compound                Strain/sex/age                    Dose and route of              Duration of       % Reduction
    (Reference)                                               administration                 the study         of hepatic
                                                                                                               vitamin A
                                                                                                                            

    Arochlor 1242           Rattus norvegicus/M,Ff/21 days       100 mg/kg in dieta           2 months             49
    (Cecil et al., 1973)

    p,p-DDT                 Rattus norvegicus/M,F/21 days        100 mg/kg in dieta           2 months             38
    (Cecil et al., 1973)

    Methoxychlor            Sprague Dawley/NRf/23 days            10 mg/kg in dietb          16 weeks               7
    (Davison & Cox, 1976)                                        100 mg/kg in diet                                 12
                                                                1000 mg/kg in diet                                 37
                                                              10 000 mg/kg in diet                                 68

    PCB                     Sprague Dawley/M/21 days             100 mg/kg in dietc           8 weeks              82
    (Innami et al., 1976)

    TCDD                    Sprague Dawley/M/NR                   10 g/kg bwf                7 days               29
    (Thunberg et al., 1979) (single oral dose)                                               14 days               39
                                                                                             28 days               59
                                                                                             56 days               67
                                                                                                                            

    Table 57 (contd).
                                                                                                                            
    Compound                Strain/sex/age                    Dose and route of              Duration of       % Reduction
    (Reference)                                               administration                 the study         of hepatic
                                                                                                               vitamin A
                                                                                                                            

    TCDD                    Sprague Dawley/M/NR                    0.1 g/kg bwg             28 days                2
    (Thunberg et al.,                                              1.0 g/kg bwg                                   27
      1980)                                                       10.0 g/kg bwg                                   65

    TCDD                    Sprague Dawley/M/2 months             15 g/kg bwg               44 days           59d  88e
    (Hakansson, 1988)                                             30 g/kg bwg                                 78d  98e
                                                                  60 g/kg bwg                                 81d  97e
                                                                 120 g/kg bwg                                 90d  99e

    Toxaphene               Sprague Dawley/M/NR                   20 mg/kg bwh               4 weeks            0
    (Thunberg et al.,
      1984)
                                                                                                                            

    a  9-12 mg vitamin A/kg diet ad libitum.
    b  33 000 IU vitamin A/kg diet ad libitum.
    c  3000 IU vitamin A/kg diet ad libitum.
    d  21 000 IU vitamin A/kg diet ad libitum.
    e  8000 IU vitamin A/kg diet ad libitum.
    f  M = male;  F = female;  NR = not reported; bw = body weight.
    g  Single oral doses.
    h  Orally twice weekly.
    


         Male Sprague Dawley rats received a single oral dose of 10 g
    TCDD/kg body weight 4 days prior to the oral administration of a
    single physiological dose of labelled vitamin A,
    (11,12-3H)retinylacetate (RA) (Hakansson & Ahlborg, 1985a). The
    distribution and elimination of the radiolabel, and the vitamin A
    content in various tissues were determined 1, 6, 12, 24, 72, and 192
    h after the administration of 3H-vitamin A. The body burden of
    radioactivity remained around 40% of the administered dose in control
    rats throughout the study, whereas in TCDD-pretreated rats the body
    burden decreased continuously from 21% at 12 h after administration to
    11% at the end of the study. More of the radiolabel was recovered in
    the kidney, testes, epididymis, and serum of TCDD-treated animals than
    in controls, when calculated as percentage of body burden, whereas
    less was recovered in the liver. Forty percent of the dose was
    eliminated via faeces and urine in TCDD-treated rats, compared to 23%
    in controls. More radioactivity was eliminated in faeces than in urine
    both in control and TCDD-pretreated animals, although urinary
    elimination was more pronounced in TCDD-pretreated than in control
    rats. It was concluded that TCDD-treated rats handled the newly
    administered dose of vitamin A in a similar way to rats deficient in
    Vitamin A (Huque, 1981; Blomhoff et al., 1982). This finding is
    remarkable since the TCDD-treated animals in this study still had
    considerable stores of hepatic vitamin A, and did not show decreased
    levels of serum vitamin A, i.e., they were not deficient in vitamin A.
    In a similarly designed study the effect of a single oral dose of 10
    g TCDD/kg body weight on the endogenous pool of vitamin A
    (radiolabelled 15-3H-retinol given 5 to 7 days prior to TCDD
    treatment) was studied in Sprague Dawley rats with stores of low liver
    vitamin A (Hakansson et al., 1986). It was demonstrated that
    endogenously stored vitamin A was rapidly depleted from the liver of
    TCDD-treated rats and was eliminated both in faeces and in urine. An
    increased distribution of the vitamin A stored in the liver to
    extrahepatic tissues was also seen in the treated rats.

         To elucidate whether dietary vitamin A would reduce TCDD
    toxicity, Hakansson (1988) fed male Sprague Dawley rats ad libitum
    from weaning throughout the experiment with diets containing 2000 (I),
    5000 (II), 8000 (III), or 21 000 (IV) IU of vitamin A/kg. A single
    oral dose of TCDD (15, 30, 60, or 120 g TCDD/kg body weight) was
    given when the rats were 8 weeks old and the animals were killed 44
    days post-treatment. With diet IV, TCDD reduced in a dose-dependent
    manner hepatic vitamin A by 59 to 90%. With diets II and III,
    reduction of hepatic vitamin A was more than 95% after dosing with 15
    and 30 g TCDD/kg, respectively. In control animals fed diet I, total
    hepatic vitamin A was less than 1 g and TCDD had no further effect at
    any dose. Serum vitamin A was dose-dependently decreased by TCDD
    treatment in dietary groups I, II, and III, whereas in dietary group
    IV TCDD increased serum vitamin A. Only with the highest TCDD dose was
    there a counteraction by dietary vitamin A on all of the above
    parameters.

         A single dose of 10 g TCDD/kg body weight to female Sprague
    Dawley rats on the day of delivery affected the vitamin A content in
    the liver and kidney of the offspring (Hakansson et al., 1987). The
    effect became clearly visible after weaning, i.e., when the dietary
    intake of vitamin A was high enough to allow for storage. At the end
    of the study (postnatal day 32), the vitamin A content in the liver
    was 225 g in control pups and 102 g in TCDD-exposed pups. The
    corresponding values for the kidney vitamin A content were 1.4 and 8.4
    g, respectively. The TCDD-induced effects on vitamin A levels in the
    liver and kidneys followed a similar time-course as the growth
    reduction, i.e., a minor effect throughout the lactation period, which
    became more pronounced post-weaning. This was in contrast to the liver
    enlargement and thymus involution in TCDD-exposed pups, which were
    most pronounced throughout lactation and tended to diminish post-
    weaning.

         Single oral doses of 0, 1, 5 and 10 g TCDD/kg body weight or a
    mixture of PCDDs and PCDFs, reconstituting the levels found in human
    milk, were given to male Sprague Dawley rats (80 g) in corn oil
    (Ahlborg et al., 1987). The mixture was given at three dose levels
    resulting in 1, 5, or 10 g TCDD/kg body weight. Four weeks after
    dosing there was a dose-dependent decrease in hepatic vitamin A in
    TCDD-treated rats; no further effect was seen in the animals treated
    with PCDD/PCDF mixtures. In contrast, mixture treatment had an
    additive effect, as compared to TCDD alone, on the increases in renal
    vitamin A and hepatic cytochrome P-450 contents, AHH activity, and
    UDPGT activity.

         Taken together these data indicate that TCDD interferes with the
    storage mechanism for vitamin A. In the liver this mechanism has been
    thoroughly investigated (Hirosawa & Yamada, 1973; Blomhoff et al.,
    1982; Olson & Gunning, 1983). As dietary vitamin A seems unable to
    counteract all toxic effects, this would imply either that the effect
    on vitamin A storage is secondary to TCDD toxicity or that the
    cellular utilization of vitamin A is affected by TCDD.

    7.5  Embryotoxicity and Reproductive Effects

         The teratogenic potential of TCDD was first demonstrated in rats
    and mice by Courtney & Moore (1971). This followed the finding that
    2,4,5-trichlorophenoxyacetic acid contaminated with 30 mg TCDD/kg was
    teratogenic in two strains of mice and one strain of rat (Courtney et
    al., 1970), leading to an increased incidence of cleft palate and
    cystic kidney in both strains of mice and cystic kidney in rats.
    Further studies have revealed that TCDD is fetotoxic rather than
    teratogenic in the rat, producing subcutaneous oedema, haemorrhages,
    and slight kidney anomalies (see section 7.5.1). In contrast, TCDD
    produces a specific teratogenic response, consisting of cleft palate
    and kidney malformations, in several strains of mice (see section
    7.5.2 and Table 58). Extra ribs, minor abnormalities in the palate,

    and cardiovascular malformations have been demonstrated in rabbits,
    monkeys, and chickens, respectively, after exposure to TCDD in utero
    (see sections 8.5.3-8.5.6). Impaired reproductive performance due to
    TCDD exposure has been demonstrated in rats and monkeys (see sections
    8.5.1 and 8.5.4).

    7.5.1  Studies on rats

         The initial studies on the embryotoxic effects of TCDD in the rat
    were performed with Charles River CD (Courtney & Moore, 1971), Sprague
    Dawley (Sparschu et al., 1971), and Wistar rats (Khera & Ruddick,
    1973). In these studies maternal toxicity was noted at doses > 0.5
    (Sparschu et al., 1971) or 1 g/kg body weight per day (Khera &
    Ruddick, 1973). Decreases in gestational survival, fetal weight, and
    postnatal survival were reported at dose levels in the range 0.125-0.5
    g TCDD/kg body weight per day on gestation days 6-15 or 5-14.
    Haemorrhages, mainly intestinal, and subcutaneous oedema were common
    findings at similar doses. The teratogenic findings were limited to
    kidney anomalies, described as unilocular cystinephrotic kidney or as
    hydronephrosis, in the CD rat at or above 0.5 g TCDD/kg per day
    (Courtney & Moore, 1971). Slightly dilated renal pelvis was reported
    in the F1 generation at the 0.01 g/kg per day dose level in a
    three-generation reproductive study of TCDD in Sprague Dawley rats
    (Murray et al., 1979). Giavini et al. (1983) found an increased
    incidence of renal anomalies in Charles River CD offspring exposed to
    2 g TCDD/kg body weight on gestation days 0 to 2, but not in Sprague
    Dawley offspring when the dam was given daily oral doses of 0, 0.125,
    0.5, and 2 g TCDD/kg body weight for 2 weeks before mating (Giavini
    et al., 1982a).

         Sparschu et al. (1971) found an increased number of resorption
    sites in Sprague Dawley rats at doses > 0.5 g TCDD/kg per day, but
    no effect on ovulation rate and preimplantation loss was reported even
    at 2 g/kg per day. TCDD exposure (0.1, 0.5, or 2.0 g/kg per day) on
    gestation days 0 to 2 had no effect on the reproductive performance of
    Sprague Dawley rats (Giavini et al., 1983). In contrast, impaired
    reproductive performance was found in Charles River CD rats after
    receiving TCDD for two weeks before mating (Giavini et al., 1982a).
    The number of resorption sites was elevated at doses of 0.5 g TCDD/kg
    per day or more, whereas the ovulation rate and preimplantation loss
    was affected only at 2 g/kg per day.

         The male reproductive ability was affected by TCDD treatment only
    at toxic doses, as judged by studies of Khera & Ruddick (1973) and
    Murray et al. (1979). Nevertheless, decreased mating frequency was
    noted in the groups, of the F0 generation that received 0.1 g
    TCDD/kg per day in the diet (Murray et al., 1979). No effect on the
    mating frequency was found by Giavini et al. (1982a, 1983).

         In a three-generation reproduction study, Murray et al. (1979)
    maintained Sprague Dawley rats on diets providing doses of 0, 0.001,

    0.01, or 0.1 g TCDD/kg body weight per day. The F0 generation
    received the diet for 90 days before mating. No toxic effects were
    observed in the F0 generation but decreased body weight and reduced
    food consumption were noted in the F1 and F2 generations at 0.01
    g/kg per day. Fertility was greatly reduced at 0.1 g/kg per day in
    the F0 generation. This group was discontinued because of the low
    number of offspring. In the F1 and F2 generations fertility was
    significantly reduced at 0.001 and 0.01 g/kg per day, respectively.
    At 0.01 g/kg per day, litter sizes were reduced, and fetal and
    neonatal survival were decreased as well as postnatal growth. Murray
    et al. (1979) concluded that doses of 0.1 and 0.01 g/kg per day
    impaired reproduction in rats. A dose of 0.001 g/kg per day had no
    effect on fertility, litter size, postnatal body weight, or neonatal
    survival and was therefore suggested to be a no-effect dose for
    reproductive lesions.

         Reevaluation of these data by Murray et al. (1979) using another
    statistical model, including pooling of the data from the four
    generations, led to the conclusion that the 0.001 g/kg per day dose
    level did affect reproduction and thus was not a no-effect level, but
    a low-effect level (Nisbet & Paxton, 1982). Kimbrough et al. (1984)
    considered that the data by Murray et al. (1979) could not be used for
    risk assessment calculations due to the great variation in fertility
    index both in controls and exposed rats.

         No embryotoxic and/or reproductive effects were found when female
    Wistar rats were exposed to 1,2,3,4-tetraCDD (50, 100, 200, 400, or
    800 g/kg body weight per day), 2,7-diCDD (250, 500, 1000 or 2000
    g/kg body weight per day); 2,3-diCDD (1000 or 2000 g/kg body weight
    per day), or 2-monoCDD (1000 or 2000 g/kg body weight per day) on
    gestation days 6-15 (Khera & Ruddick, 1973). The maturation process of
    the lung was not affected in Sherman rats exposed to 2,7-diCDD (40
    g/kg per day on gestation days 7-15) (Kimbrough et al., 1974).

    7.5.2  Studies on mice

         TCDD-induced embryo mortality in NMRI mice was significantly
    increased at doses of 4.5 and 9 g TCDD/kg body weight per day when
    given on gestation days 6-15, but no embryotoxic effect was observed
    when daily doses of 9 g/kg body weight were given on gestation days
    9-13 (Neubert & Dillman, 1982). The number of resorptions on day 13 of
    gestation in NMRI mice was increased, as compared to controls, when 25
    g TCDD/kg body weight was given, divided into five daily doses on
    days 7-11 of gestation, but no effect was observed when the same dose
    was given as single ip injections on days 7 or 10 of gestation (Nau &
    Bass, 1981). A single ip dose of 30 g TCDD/kg body weight on
    gestation day 11 had no effect on the fetal mortality in C57Bl/6 mice
    on days 12, 13, or 14 of gestation (Weber & Birnbaum, 1985). Decreased
    pre- and postnatal survival rates and retarded postnatal development
    were observed in NMRI mice given four oral doses of 12.5 g TCDD/kg
    body weight on gestation days 14-17 (Nau et al., 1986). The cumulative

    mortality was 45%, 68.5%, and 75% in exposed offspring on postnatal
    days 1, 14, and 22, respectively, as compared to 6% in controls on day
    22.

         Table 58 summarizes the early studies on the teratogenic effects
    of TCDD in various strains of mice. These studies (Courtney & Moore,
    1971; Neubert & Dillman, 1972; Courtney, 1976; Smith et al., 1976),
    revealed that TCDD is a specific teratogen in mice causing increased
    frequencies of kidney anomalies and cleft palate at doses well below
    those which result in fetal mortality and maternal toxicity.

         The TCDD-induced kidney anomaly is morphologically described as
    a progressive hydronephrosis, preferentially occurring in the right
    kidney, and never accompanied by hydroureter or abnormal nephron
    development (Courtney & Moore, 1971; Moore et al., 1973; Birnbaum et
    al., 1985; Weber et al., 1985).

         The fetal kidney seems to be more susceptible to TCDD exposure
    than the developing palate (Courtney & Moore, 1971; Moore et al.,
    1973; Birnbaum et al., 1986; Weber et al.,1984). The incidence of
    kidney anomalies after a single dose of 1 g TCDD/kg body weight on
    gestation day 10 in C57Bl/6 mice was 34.3%. If the same dose was
    divided and given on gestation days 10-13, the incidence of kidney
    anomalies was 58.9%. The incidences of cleft palate in the same
    studies were 0 and 1.9%, respectively (Moore et al., 1973). In
    contrast, in the CF-1 strain of mice, cleft palate was a more
    sensitive parameter than hydronephrosis, occurring at 1 and 3 g
    TCDD/kg body weight per day on gestation days 6 to 15, respectively
    (Smith et al., 1976). It would appear that 3 g TCDD/kg per day on
    gestation days 10-13 is close to a threshold dose for cleft palate
    induction in C57Bl/6 mice (Moore et al., 1973; Birnbaum et al., 1986).
    The maximum increase in cleft palate incidence is produced if TCDD is
    administered on any individual day from gestation day 8-10 in C57Bl/6
    mice (Pratt et al., 1984) or on gestation day 10-11 in NMRI mice (Nau
    & Bass, 1981; Neubert & Dillman, 1982; Krowke, 1986). The cleft palate
    incidence in C57Bl/6 mice was more pronounced (36%) when TCDD was
    given as a single oral dose of 12 g/kg body weight on gestation day
    11 than if 3 g/kg body weight per day was given on gestation days
    10-13 (Birnbaum et al., 1985). No difference was seen in the incidence
    of kidney anomalies with the same treatment. TCDD is not a potent
    inducer of cleft palate when given on day 13 of gestation or later
    (Neubert & Dillman, 1982; Pratt et al., 1984). Examination of cryostat
    sections taken from C57Bl/6 embryos during the time of palatal
    elevation and fusion demonstrated that TCDD does not interfere with
    growth, elevation, or initial contact of the palatal shelves but does
    interfere with the firm adhesion and/or degeneration of the medial
    epithelial cells, i.e., programmed epithelial cell death does not
    appear to occur in the medial epithelium in embryos exposed to TCDD
    (Pratt et al., 1984). The cleft palates that were observed were
    complete clefts of the entire hard and soft palate and no clefts of
    the primary palate were observed (Pratt et al., 1984).



    
    Table 58. Early studies on embryotoxic effects of TCDD in mice
                                                                                                                            

    Strain/sexb       Route/vehicle/dose       Treatment/          Parental toxicity             Embryotoxic effects
    (reference)         (g/kg body weight       observation
                              day)               (days)a
                                                                                                                            
    CD-1/F              subcutaneous/            6-15 / 18           No effect                     Increased incidence of kidney
    (Courtney &         DMSO/1.0, 3.0                                                              anomalies at doses > 1 g/kg
      Moore, 1971)                                                                                 per day.

    BDA/2J              subcutaneous/            6-15 / 17           Increased relative            Increased incidence of cleft
    (Courtney &         DMSO/3.0                                     liver weight                  palate and kidney anomalies.
      Moore, 1971)

    C56BL/6J            subcutaneous/            6-15 / 17           Increased relative            Increased incidence of cleft
    (Courtney &         DMSO/3.0                                     liver weight                  palate and kidney anomalies.
      Moore, 1971)

    NMRI/F              oral/rape seed           6-15 / 18           None reported                 Increased number of resorptions
    (Neubert &          oil/0.3, 3.0, 4.5,                                                         at 9 g/kg per day. Increased
      Dillman, 1972)    9.0                                                                        incidence of cleft palate at
                                                                                                   doses > 3 g/kg per day.

    C56BL/6J            oral/corn oil:acetone    10-13 / 18          None reported                 Increased incidence of cleft
    (Neubert &          (9:1)/1.0, 3.0                                                             palate at doses > 3 g/kg
      Dillman, 1972)                                                                               per day increased incidence
                                                                                                   of kidney anomalies at doses
                                                                                                   > 1 g/kg per day.
                                                                                                                            

    Table 58 (contd).
                                                                                                                            

    Strain/sexb         Route/vehicle/dose       Treatment/          Parental toxicity             Embryotoxic effects
    (reference)         (g/kg body weight       observation
                              day)               (days)a
                                                                                                                            

    CD-1/F              oral/corn oil:anisole    7-16 / 18           Increased maternal            Increased fetal mortality at all
    (Courtney, 1976)    (95:5)/25, 50, 100,                          relative liver weight         doses. Increased incidence of
                        200, 400                                     at 25 and 50 g/kg per        kidney anomalies and cleft palate
                                                                     day. Marked oedema and        at doses > 25 g/kg per day
                                                                     vaginal bleeding at           and 50 g/kg per day,
                                                                     doses > 200 g/kg per        respectively. Increased incidence
                                                                     day.                          of club foot was found in the
                                                                                                   high-dose group. Hydrocephalus and
                                                                                                   open eyes were seen occasionally.

    CF-1                oral/corn oil:acetone    6-15 / 18           Increased maternal            Increased number of resorptions
    (Smith et al.,      (98:2)/0.001, 0.01,                          relative liver weight         at 1 g/kg per day. Increased
      1976)             0.1, 1.0, 3.0                                at 3 g/kg per day.           incidence of cleft palate at
                                                                                                   doses > 1 g/kg per day.
                                                                                                   Dilated renal pelvis occurred at
                                                                                                   3 g/kg per day.

                                                                                                                            

    a  First day of gestation designated day zero.
    b  F = female.
    


         Poland & Glover (1980) reported that in nine out of ten inbred
    strains of mice, the susceptibility to cleft palate formation produced
    by TCDD (30 g TCDD/kg body weight sc on day 10 of pregnancy) followed
    the distribution of the Ah locus within that strain. The five strains
    with a low affinity TCDD receptor in the liver (DBA/2, RI, AKR, SWR,
    and 129) developed cleft palate at an incidence of 0 to 3% while four
    of the five strains with a high affinity TCDD receptor in the liver
    (C57Bl/6, A, BALB/cBy, and SEC) developed cleft palate at an incidence
    of 54 to 95%. The only strain with a high affinity hepatic TCDD
    receptor that did not develop cleft palate was CBA. Mid-gestational
    mice embryos from C57Bl/6 exhibit high levels of the TCDD receptor in
    the maxillary processes and secondary palatal shelves whereas no
    specific binding of TCDD could be demonstrated in the AKR strain
    (Dencker & Pratt, 1981). Evidence that TCDD interferes in embryonic
    development by directly interacting with embryonic cells rather than
    being secondary to maternal effects was presented by D'Argy et al.,
    (1984) in a study where mouse blastocysts were transplanted between
    NMRI (sensitive) and DBA (non-sensitive) dams on gestation day 3. TCDD
    treatment (30 g/kg body weight) on gestation day 10 resulted in 75 to
    100% incidence of cleft palate among NMRI fetuses whether they
    remained in their own dams or as aliens in DBA dams. Also, none of the
    DBA fetuses developed cleft palate whether or not they remained in
    their own dams or as aliens in NMRI dams.

         Attempts to modify and further characterize the TCDD-induced
    cleft palate formation have been performed with several interacting
    substances. The non-teratogenic -naphthoflavone (N) enhanced the
    TCDD-induced fetal mortality and the increase in the incidence of
    cleft palate in C57Bl/6 and NMRI mice when N was administered
    simultaneously or 8 h before TCDD, but not 24 h before or after
    (Hassoun & Dencker, 1982). TCDD was given as a single ip dose of 25 or
    16 mg/kg body weight to C57Bl/6 and NMRI mice, respectively, on
    gestation days 10, 11, 12, or 13. TCDD-induced cleft palate incidence
    was enhanced by 2,3,4,5,3',4'-hexachlorobiphenyl, but not
    2,4,5,2',4',5'-hexachlorobiphenyl, in C57Bl/6 mice, although none of
    the isomers alone were teratogenic at the doses used (Birnbaum et al.,
    1985). Neither of the hexachlorobiphenyls used affected the incidence
    of renal anomalies, as compared to TCDD alone.

         A dose-related enhancement of the TCDD-induced incidence of cleft
    palate was found in C57Bl/6 mice exposed to either triiodothyronine or
    thyroxine, as compared to TCDD alone (Lamb et al., 1986). The
    hydrocortisone-induced cleft palate response was enhanced by
    simultaneous administration of TCDD in C57Bl/6 mice (Birnbaum et al.,
    1986). Co-administration of TCDD and TCDF to C57Bl/6 mice on gestation
    day 10 resulted in a cleft palate incidence compatible with an
    additive toxicity model in which TCDF contributes to the toxicity of
    TCDD in a weight ratio of 1:30 (Weber et al., 1985). Also 1,2,3,7,8-
    pentaCDD and 1,2,3,4,7,8-hexaCDD had additive effects on the
    TCDD-induced cleft palate incidence in NMRI mice (Krowke, 1986). No

    clearly dose-related teratogenic effects were observed in C57Bl/6 mice
    exposed to oral doses of 1,2,3,4- tetraCDD (100, 250, 500, or 1000
    mg/kg per day), octaCDD (5 or 20 mg/kg per day) or to a mixture of 40%
    2,7-diCDD and 60% 2,3,7-triCDD (100 or 200 mg/kg per day) on gestation
    days 7-16 (Courtney, 1976).

    7.5.3 Studies on rabbits

         New Zealand rabbits were administered TCDD by gavage in doses of
    0, 0.1, 0.25, 0.5, and 1 g/kg per day on days 6-15 of gestation and
    the fetuses were examined on day 28 of gestation (Giavani et al.,
    1982b). Above 0.25 g/kg per day, decreased maternal weight gain and
    unspecified signs of maternal toxicity were reported. At doses of 0.5
    and 1 g/kg per day, there were 2/15 and 4/10 maternal deaths,
    respectively. An increase in abortion and resorption rates occurred at
    doses above 0.25 g/kg per day, with no live fetuses detected in the
    1 g per kg per day dose group. There was a significant increase in
    extra ribs, compared to a level of 33.3% in the controls to 82, 66.6,
    and 82%. In the 0.1, 0.25, and 0.5 g/kg per day dose groups, 82,
    66.6, and 82% extra ribs were noted. There were no increases in
    specific soft tissue anomalies.

    7.5.4  Studies on monkeys

         The effect of exposure to TCDD in the diet, before mating and
    throughout gestation, on the reproductive performance and production
    of progeny of healthy, fertile rhesus monkeys (Macaca mulatta) was
    studied by Allen and co-workers (Allen et al., 1979a,b; Barsotti et
    al., 1979; Schantz et al., 1979). At a level of 500 ng TCDD/kg diet
    (11 ng/kg body weight per day), there were no effects on the length,
    intensity, or duration of the menstrual cycle, but decreases in serum
    estradiol and progesterone levels were observed (Barsotti et al.,
    1979). Mating with control males at the end of the sixth month
    resulted in three pregnancies, out of which two resulted in abortion,
    and three animals failed to conceive. The remaining TCDD-treated
    female was not bred due to toxic symptoms. The two females that
    survived the study were returned to a control diet and later gave
    birth to well developed infants. After 7 months on the 50 ng TCDD/kg
    diet (1.5 ng/kg body weight per day), the reproductive outcome was:
    four abortions, one stillbirth, two failures to conceive and two
    normal births (Schantz et al., 1979).

         All controls conceived and gave birth to normal infants (Barsotti
    et al., 1979; Schantz et al., 1979).

         McNulty (1984) demonstrated that rhesus monkeys (Macaca mulatta)
    receiving 1 g TCDD/kg body weight either as a single dose on
    gestation days 25, 30, 35, or 40, or divided into nine doses between
    gestation days 20-40, failed to give birth normally. Of 16 pregnancies
    13 resulted in abortions. Two of the three live fetuses, obtained by
    Caesarean section on day 145 of gestation, showed minor abnormalities

    in the palate. Maternal toxicity, manifest in 8 out of 16 females,
    appeared only after a period of 44-111 days after abortion. From the
    results obtained it was not possible to conclude whether fetal death
    was a direct effect of TCDD on the fetus or placenta or an indirect
    effect through maternal toxicity.

    7.5.5  Studies on chickens

         Treatment of fertile White Leghorn chicken eggs, on day 0 of
    development, with single doses of TCDD, ranging from 0.009 to 77.5
    pmol/egg, resulted in a dose-related increase in the following types
    of cardiovascular malformations: ventricular septal defects, aortic
    arch anomalies, aortic arch anomaly plus ventricular septal defect,
    and conotruncal malformation (Cheung et al., 1981). The dose producing
    cardiovascular malformations in 50% of the embryos was about 1 pmole
    TCDD/egg.

         A dose-dependent decrease in hatchability and increased
    incidences of beak, brain, and leg malformations were found in the
    embryos when fertile White Leghorn eggs were injected with toxic fat
    material containing 0.9, 1.8, or 4.5 ng of a mixture of dioxins (14%
    diCDDs, 1% triCDDs, 38 to 45% tetraCDDs, 13% pentaCDDs, 14% hexaCDDs,
    12% heptaCDDs, 8% octaCDD) (Flick et al., 1973).

    7.6  Mutagenicity and Related End-Points

    7.6.1  Mutagenicity

    7.6.1.1  Studies on bacteria

         Results from bacterial mutagenicity tests with TCDD are
    conflicting.

         In studies by Hussain et al. (1972) and Seiler (1973), a positive
    response was reported in the Salmonella typhimurium strain TA 1532
    without metabolic activation in a plate test after preincubation of
    bacteria in medium containing TCDD and also in a spot test. A
    mutagenic response was also obtained with Escherichia coli Sd-4,
    measuring reversion to streptomycin independence (Hussain et al.,
    1972). OctaCDD was not mutagenic to various strains of Salmonella
    typhimurium (Seiler, 1973) without metabolic activation. More recent
    publications, however, do not report any mutagenic effect of TCDD in
    the Ames' Salmonella plate incorporation assay using strains TA
    1530, TA 1532, TA 1535, TA 1537, TA 1538, TA 98, or TA 100 in the
    presence or absence of metabolic activation systems from rat and
    Syrian hamster liver (Gilbert et al., 1980; Geiger & Neal, 1981;
    Mortelmans, 1984). In these studies the earlier reported positive
    tester strain TA 1532 was either replaced by strain TA 1537 (Geiger &
    Neal, 1981; Mortelmans, 1984) or was tested in addition to TA 1537
    (Gilbert et al., 1980). Strain TA 1537 has been derived from TA 1532

    and is more sensitive, due to its improved uptake of large molecules.
    TCDD was tested in the dose range 0.2-2000 g/plate. Due to the
    limited solubility of TCDD, a maximal dose in the Salmonella system
    was reported to be 20 g TCDD/plate (Geiger & Neal, 1981).

    7.6.1.2  Studies on eukaryotic cells

         Bronzetti et al. (1983) reported a mutagenic response of TCDD in
    yeast in an in vitro suspension test and a host mediated assay. In
    both assays Saccharomyces cerevisiae strain D7 was used. Positive
    responses were obtained in vitro in the presence of metabolic
    activation at doses of TCDD up to 10 g TCDD/ml, and in the
    host-mediated assay after treatment of mice with a single dose of TCDD
    (25 g/kg body weight).

         In L5178Y mouse lymphoma cells, TCDD induced mutations in a
    dose-dependent manner at doses of 0.05-0.5 g TCDD/ml; survival of
    cells at the highest concentration was at least 75% (Rogers et al.,
    1982).

    7.6.1.3  In vivo studies

         A dominant lethal test on male Wistar rats has been reported by
    Khera & Ruddick (1973). The rats were given 4, 8, or 12 g TCDD/kg
    body weight per day orally for seven consecutive days, after which
    seven sequential mating trials, 5 days at a time, were conducted in
    the surviving males. Nine days after separation of females from the
    males, the females were killed. The highest dose was lethal for all
    males exposed (20/20); 8 g/kg per day killed 11/20 males, and 4 g/kg
    per day was lethal for 2/20. All animals in the control group
    survived.

         The results did not indicate a dominant lethal effect during the
    35 days post-treatment period, corresponding to the postmeiotic stages
    of spermatogenesis.

    7.6.2 Interaction with nucleic acids

         Poland & Glover (1979) found that very little TCDD bound to rat
    liver nucleic acids after treatment with 3H-TCDD in vivo; the
    maximum covalently bound TCDD was calculated to be 6 and 12 pmol per
    mol of nucleotide residues from DNA and RNA, respectively. After iv
    injection of 3H-TCDD in rats, the radioactivity taken up by the
    liver cytosol decreased at the same rate as the radioactivity in the
    nuclear fraction increased. The radioactivity in the nuclei was at a
    maximum 2 h after injection (Carlstedt-Duke et al., 1982). Guenthner
    et al. (1979) demonstrated in vitro metabolism of TCDD to reactive
    intermediates that bound covalently to cellular macromolecules,
    principally to proteins. However, no isomer-specific methods were used
    for analysis of metabolites in this study.

         Liver slices from Sprague Dawley rats treated with 5 g TCDD
    incorporated twice the level of thymidine into nuclear DNA 10 days
    post-treatment than did controls (Conaway & Matsumura, 1975).
    Christian & Peterson (1983a) found no effect on the in vivo
    incorporation of thymidine into liver DNA 35 to 36 h after the
    administration of 10 g TCDD/kg body weight to Sprague Dawley rats.
    DNA synthesis in Porton rats, stimulated by 70% hepatectomy and
    measured as the 1 h in vivo incorporation of thymidine, was not
    affected by treatment with 10 or 200 g TCDD/kg body weight 0, 24, or
    72 h before the hepatectomy was performed (Greig et al., 1974). In
    contrast, Dickins et al. (1981) found an 8-to 10-fold increase in DNA
    synthesis in response to the proliferation caused by a 1/3 hepatectomy
    in Sprague Dawley rats given 5 g TCDD/kg body weight 5 days prior to
    the hepatectomy. In this study thymidine incorporation into DNA was
    measured in vitro at various times after hepatectomy. The
    TCDD-induced increase was most pronounced 24 to 32 h after the
    hepatectomy. The somewhat conflicting results may be due to
    differences in the in vivo and in vitro incorporation of
    thymidine, as well as to differences in the time points studied.
    According to Dickins et al. (1981), the discrepancy in proliferative
    DNA synthesis could be due to the degree of hepatectomy. Thus 70%
    hepatectomy would by itself enhance DNA synthesis to near maximum
    level, making it difficult to measure any effect of TCDD under the
    experimental conditions. This suggestion was confirmed by Christian &
    Peterson (1983a), who compared the effect of TCDD on proliferative DNA
    synthesis after 1/3 and 2/3 hepatectomy. This study also revealed that
    the effect could be seen only when a certain amount of time, namely 5
    to 10 days, elapsed between TCDD administration and hepatectomy.

         The transfectivity of bacteriophage Q/RNA was evaluated after
    treatment in vitro with TCDD. No effect was noticed in the tested
    dose range (0.2-4 g TCDD/ml) (Kondorosi et al., 1973).

         The effect of TCDD on the repair of DNA damage induced by
    2-aminofluorene (AF) and 2-acetylaminofluorene (AAF) in primary
    hepatocytes from B6 and D2 mice has been investigated (Moller et al.,
    1984). Pretreatment in vivo with TCDD (50 mmol/litre) resulted in
    a slight increase in DNA damage (measured by the alkaline elution
    technique) following incubation with either AF or AAF for 60 min,
    suggesting induction of aromatic amine activating enzymes.

    7.6.3  Cytogenetic effects

         Green & Mooreland (1975) and Loprieno et al. (1982a) did not
    observe any induction of chromosomal aberrations in rats administered
    TCDD intraperitoneally or by gavage (5 to 20 g/kg body weight). Later
    Green et al. (1977) showed a significant increase in the induction of
    chromosomal abnormalities in bone marrow cells of male rats at doses
    of 2 and 4 g TCDD/kg body weight and in females at 4 g/kg body
    weight. In a study by Meyne et al. (1985) C57Bl/6 or DBA/2 mice (with

    high- and low-affinity TCDD receptors, respectively) were ip injected
    at doses 0, 50, 100, and 150 g/kg body weight. There was no increase
    in the frequency of chromosomal aberrations in bone marrow cells of
    the TCDD-treated mice of either strain 8, 16, or 24 h after treatment.
    All doses were high enough to induce hepatotoxic damage in C57Bl/6
    mice. In male and female CD-1 mice, however, a weak but significant
    increase in chromosomal aberrations was obtained 96 h post-treatment
    (ip 10 g TCDD/kg body weight) (Loprieno et al., 1982b).

         Lamb et al. (1981) evaluated the frequency of sister chromatid
    exchange (SCE) in bone marrow cells of C57Bl/6 mice given single ip
    injections of mixtures of chlorinated phenoxy acids that contained
    0.16, 1.2, or 2.4 g TCDD/kg body weight. The mean SCE frequency in
    treated and in control mice did not differ significantly. Mice fed
    diets containing the same daily doses as above for 4 to 8 weeks did
    not show any increase in SCE frequencies.

         At doses (50, 100, and 150 g per kg body weight) hepatotoxic to
    C57Bl/6 mice, there was no significant increase of SCEs in C57Bl/6 or
    DBA/2 mice 18 h after ip injection of TCDD (Meyne et al., 1985). Meyne
    et al. (1985) also performed a micronucleus test under the same
    conditions as above. Mice, killed 24 or 48 h after treatment, did not
    show any increase of micronuclei in polychromatic erythrocytes of bone
    marrow in either strain.

    7.6.4 Cell transformation

         Single treatments (11 concentrations in the range 0 to 5
    mol/litre) of mouse embryo fibroblast (C3H/10T1/2) tissue cultures
    with TCDD did not transform or initiate the process of transformation
    in cultures subsequently exposed to
    12-O-tetradecanoylphorbol-13-acetate (Abernethy et al., 1985).
    Continued treatment of these cells with low concentrations (> 4
    pmol/litre) of TCDD enhanced the production of foci in cultures
    pretreated with N-methyl-N'-nitro-N-nitrosoguanidine. Maximal
    enhancement occurred at 40 pmol/litre. Higher doses, 120 to 4000
    pmol/litre, did not further increase the incidence of foci production.
    Promotion of transformation is thus the predominant effect of TCDD in
    the C3H/10T1/2 cell-transformation system.

    7.7 Carcinogenicity

    7.7.1  Long-term animal studies on single compounds

         Several studies on the carcinogenicity of TCDD and related
    compounds have been performed.

         The data from studies using oral exposure are summarized in Table
    59. Van Miller et al. (1977) exposed male Sprague Dawley rats to
    various dietary levels of TCDD ranging from 0.001 g/kg and 1 g/kg to
    1 g/kg for 78 weeks. Pronounced mortality was observed at higher
    doses. Neoplastic changes in different organs were noted in a number
    of rats that died. At 95 weeks, the small number of surviving animals
    were killed. At dietary levels of 5, 50, and 500 ppt TCDD (ng/kg
    feed), a variety of tumours were noted, but no particular trend
    emerged. However, at a level of 5 g/kg feed, four squamous cell
    tumours of the lung, four neoplastic nodules (hyperplastic nodules),
    and two cholangiocarcinomas of the liver were found in seven rats.

         Kociba et al. (1978) fed groups of 50 male and female Sprague
    Dawley rats 0.1, 0.01, and 0.001 g TCDD/kg body weight for 2 years.
    86 male and 86 female control rats received the vehicle only. The
    doses corresponded to 2193, 208, and 22 ng TCDD/kg diet. A variety of
    tumours were found in the control and experimental groups. Tumours
    caused by the ingestion of TCDD were confined to the liver, lungs,
    hard palate/nasal turbinates, and tongue. In the female rats that had
    received doses of 0.1 and 0.01 g/kg body weight, a statistically
    significant increase of neoplastic nodules (hyperplastic nodules,
    hepatomas) of the liver was noted, and in the rats that had received
    0.1 mg TCDD/kg body weight there was a statistically significant
    increase of hepatocellular carcinomas. Epithelial tumours along the
    respiratory tract, tongue, and hard palate consisted of well
    differentiated squamous cell carcinomas. There was an increased
    incidence, compared with the controls, of squamous cell carcinomas of
    the hard palate and nasal turbinate in both male and female rats
    receiving 0.1 g TCDD/kg body weight, while the incidence of squamous
    cell carcinoma of the lungs at this dose showed an increase only in
    the females. The authors also noted a decreased incidence of tumours
    of the pituitary gland, uterus, mammary glands, pancreas, and adrenal
    glands in the treated groups, possibly secondary to an effect on the
    hormonal functions of different glands. This decrease was in some
    instances statistically significant.

         Two further studies on the carcinogenicity of TCDD are available
    (NIH, 1982a; NIH, 1982b). The TCDD used in these studies was reported
    to be 99.4% pure, based on gas chromatographic analysis.

         In two gavage studies both Osborne-Mendel rats and B6C3F1 mice
    were used (NIH, 1982a). All animals were about 6 weeks old. Dosages,
    duration, and outcome are summarized in Table 59. The statistical
    analysis performed was similar to that in the dermal study (NIH,
    1982b). Mean body weights of the high-dose groups of rats were lower
    than those of the corresponding controls after week 55 and 45 for
    males and females, respectively, but no other clinical signs were
    observed. No such dose-related depression in mean body weight gain was
    observed in mice when compared to the vehicle-control groups.



    
    Table 59. Carcinogenicity bioassays of PCDDs after oral administration
                                                                                                                            
    Compound                 Exposure: route, dose, frequency   Species/strain/sex       Tumour type and incidence
    (Reference)              and duration-treatment/test
                                                                                                                            
    2,3,7,8-TCDD             Oral (diet), 0.0, 0.001, 0.005,    Rat/Sprague Dawley/M     all tumours: 0/10 at 0.0,
    (Van Miller et al.,      0.05, 0.5, 1.0, 5.0 g/kg.                                  0/10 at 0.001, 5/10 at 0.005,
      1977)                  78/95 weeks                                                 3/10 at 0.05, 4/10 at 0.5, 4/10
                                                                                         at 1.0, and 7/10 at 5.0 g/kg

    2,3,7,8-TCDD             Oral (diet), 0.0, 0.001, 0.01,     Rat/Sprague Dawley/M     squamous cell carcinoma hard
    (Kociba et al., 1978)    0.1 g/kg body weight per day.                              palate: 4/50 at 0.1 g/kg per day;
                             105/105 weeks                                               squamous cell carcinoma tongue:
                                                                                         1/50 at 0.001 and 0.01, 3/50 at
                                                                                         0.1 g/kg per day; adenoma of
                                                                                         adrenal cortex: 2/5 at 0.01 and
                                                                                         5/50 at 0.1 mg/kg per day

                                                                Rat/Sprague Dawley/F     hepatocellular carcinoma: 0/86
                                                                                         at 0.0, 0/50 at 0.001, 2/50 at
                                                                                         0.01, and 11/49 at 0.1 g/kg per
                                                                                         day; squamous cell carcinoma of
                                                                                         tongue: 1/50 at 0.01 and 4/49 at
                                                                                         0.1 g/kg per day, squamous cell
                                                                                         carcinoma of lung: 7/49 at 0.1
                                                                                         g/kg per day.

    2,3,7,8-TCDD             Oral (gavage corn oil:acetone,     Rats/Osborne-Mendel/M    follicular cell adenomas or
    (NIH, 1982a)             9:1), 0.0, 0.1, 0.05, 0.5                                   carcinoma of thyroid: 1/69 at
                             g/kg body weight per week.                                 0.0, 5/48 at 0.10, 8/50 at 0.05,
                             104/105-107 weeks                                           and 11/50 at 0.5 g/kg per week
                                                                                                                            

    Table 59 (contd - 2).
                                                                                                                            

    Compound                 Exposure: route, dose, frequency   Species/strain/sex       Tumour type and incidence
    (Reference)              and duration-treatment/test
                                                                                                                            

                                                                Rats/Osborne-Mendel/F    follicular cell adenomas or carcinoma
                                                                                         of thyroid: 3/73 at 0.0, 2/45 at 0.1,
                                                                                         1/49 at 0.05, and 6/47 at 0.5 g/kg
                                                                                         per week; neoplastic nodules or
                                                                                         hepatocellular carcinoma: 5/75 at
                                                                                         0.0, 1/49 at 0.1, 3/50 at 0.05, and
                                                                                         14/49 at 0.5 g/kg per week

    2,3,7,8-TCDD             Oral (gavage corn oil:acetone,     Mice/B6C3F18/M           hepatocellular carcinoma: 8/73
    (NIH, 1982a)             9:1), Males 0.0, 0.01, 0.05,                                at 0.0, 9/49 at 0.01, 8/49 at
                             0.5 g/kg body weight per                                   0.05, and 17/50 at 0.5 g/kg per
                             week. Females 0.0, 0.04, 0.2,                               week
                             2.0 g/kg body weight per
                             week, 104/105 weeks                Mice/B6C3F18/F           hepatocellular carcinoma: 1/73
                                                                                         at 0.0, 2/50 at 0.04, 2/48 at
                                                                                         0.2, and 6/47 at 2.0 g/kg per
                                                                                         week; follicular cell adenomas
                                                                                         of thyroid: 0/69 at 0.0, 3/50
                                                                                         at 0.04, 1/47 at 0.2, and 5/46
                                                                                         at 2.0 g/kg per week

    1,2,3,6,7,8/             Oral (gavage corn oil:acetone,     Rats/Osborne-Mendel/M    liver neoplastic nodules or
    1,2,3,7,8,9-             9:1), 0.0, 1.25, 2.5, 5.0 g/kg                             hepatocellular carcinoma: 0/74
    hexaCDD (1:2)            body weight per week,                                       at 0.0, 0/49 at 1.25, 1/50 at
    (NIH, 1980b)             104/105 weeks                                               2.5, and 4/48 at 5.0 g/kg per
                                                                                         week

                                                                Rats/Osborne-Mendel/F    liver neoplastic nodules or
                                                                                         hepatocellular carcinoma: 5/75
                                                                                         at 0.0, 10/50 at 1.25, 12/50 at
                                                                                         2.5, and 30/50 at 5.0 g/kg per
                                                                                         week
                                                                                                                            

    Table 59 (contd - 3).
                                                                                                                            
    Compound                 Exposure: route, dose, frequency   Species/strain/sex       Tumour type and incidence
    (Reference)              and duration-treatment/test
                                                                                                                            
    1,2,3,6,7,8-/            Oral (gavage corn oil:acetone,     Mice(B6C3F1)/M           hepatocellular adenomas or carcinomas:
    1,2,3,7,8,9-             9:1), Males 0.0, 1.25, 2.5,                                 15/73 at 0.0, 14/50 at 1.25, 14.49 at
    hexaCDD (1.2)            5.0 g/kg body weight per week                              2.5, and 24/48 at 5.0 g/kg per
    (NIH, 1980b)             Females 0.0, 2.5, 5.0, 10.0 g/                             week
                             kg body weight per week,
                             104/105-108 weeks                  Mice(B6C3F1)/F           hepatocellular adenomas or carcinomas:
                                                                                         3/73 at 0.0, 4/48 at 2.5, 6.47 at 5.0,
                                                                                         and 10/47 at 10.0 g/kg per week

    Dibenzo-p-dioxin         Oral (diet), 0, 5000, 10 000       Mice/B6C3F1/M            hepatocellular carcinoma: 4/49 at 0,
    (NCI, 1977)              g/kg diet, 87-90/91-97 weeks                               7/50 at 5000, and 3/48 at 10 000
                                                                                         g/kg diet; hepatocellular adenomas:
                                                                                         4/49 at 0, 1/50 at 5000, and 2/48 at
                                                                                         10,000 mg/kg diet; malignant
                                                                                         tumours: 5/49 at 0, 11/50 at 5000,
                                                                                         and 8/50 at 10 000 mg/kg diet

                                                                Mice/B6C3F1/F            malignant tumours: 8/50 at 0,
                                                                                         9/49 at 5000, and 3/39 at 10,000
                                                                                         mg/kg diet; hepatocellular
                                                                                         carcinoma: 1/47 at 5000 mg/kg
                                                                                         diet

    2,7-diCDD                Oral (diet), 0, 5000, 10 000       Rats/Osborne-Mendel/M    malignant tumours: 5/33 at 0,
    (NCI, 1979)              mg/kg diet, 110/110-117 weeks.                              7/34 at 5000, and 4/33 at 10 000
                                                                                         mg/kg diet; hepatocellular
                                                                                         adenoma: 1/33 at 0; hepatocellular
                                                                                         carcinoma: 1/33 at 10,000 mg/kg diet

                                                                Rats/Osborne-Mendel/F    malignant tumours: 5/31 at 0;
                                                                                         4/33 at 5000, and 5/30 at
                                                                                         10 000 mg/kg diet

                                                                                                                           
    


         In the male rats, increased incidences of follicular cell
    adenomas or carcinomas in the thyroid were dose related and were
    significantly higher (P < 0.001) in the high-dose group than in the
    vehicle controls (1%, 10%, 16%, and 22%). In the female rats, an
    increase (though not statistically significant) was seen only in the
    high-dose group (4%, 4%, 2%, and 13%). The incidence of neoplastic
    nodules of the liver in the high-dose group of female rats was
    significantly (P < 0.006) higher than that in the vehicle-control
    group (7%, 2%, 6%, and 28%).

         In male and female mice, incidences of hepatocellular carcinomas
    were dose related and, in the high-dose groups, were significantly (P
    < 0.002 and 0.014, respectively) higher than those in the
    corresponding vehicle-control groups (males: 11%, 18%, 16%, and 34%;
    females: 1%, 4%, 4%, and 13%).

         Follicular cell adenomas in the thyroid occurred at dose-related
    incidences in female mice, and were significantly (P < 0.009) higher
    in the high-dose groups than those in the vehicle controls (0%, 6%,
    2%, and 11%). In conclusion, under the conditions of this bioassay,
    TCDD was carcinogenic for Osborne-Mendel rats, inducing follicular
    cell thyroid adenomas in males and neoplastic nodules of the liver in
    females. TCDD was also carcinogenic for B6C3F1 mice, inducing
    hepatocellular carcinomas in males and females and follicular cell
    thyroid adenomas in females.

         Toth et al. (1979) administered TCDD orally by gavage to groups
    of 45 male Swiss/H/Riop mice at doses of 0, 0.007, 0.7, and 7 g/kg
    body weight once a week for one year, and the animals were followed
    for their lifetime. Liver tumours were found at 18%, 29%, 48%, and
    30%, respectively. The tumour incidence at 0.7 g/kg was significantly
    higher when compared to controls (P < 0.01), while the increase at
    the highest dose level (7 g/kg) was not statistically significant (P
    = 0.11). The latter finding may be due to a much reduced average
    survival in comparison with the control group (average life span 424
    and 588 days, respectively).

         In a study using dermal application of TCDD (NIH, 1982b), male
    and female Swiss-Webster mice were about 6 weeks old at the beginning
    of the bioassay. The one-tailed Fisher exact test was used to compare
    the tumour incidence of a control group with that of a group of dosed
    animals. Mean body weights of dosed animals were essentially the same
    as those of the corresponding vehicle-control groups, but less than
    those of the untreated controls, for males throughout the study and
    for females during the first 80 weeks. The incidence of fibrosarcoma
    in the integumentary system of female mice treated with TCDD or TCDD
    and dimethylbenzathraline (DMBA) was significantly higher than that of
    the controls (P < 0.007 and P < 0.010, respectively). An increase in
    the same tumour type, although not statistically significant (P =
    0.084), was also observed in the male mice (7% and 21% for the control

    and TCDD-treated groups, respectively). In conclusion, under the
    conditions of this bioassay, TCDD was carcinogenic for female
    Swiss-Webster mice, causing fibrosarcomas in the integumentary system.
    However, the study has been criticized in several areas, namely, a
    maximal tolerated dose (MTD) was not achieved, especially in male
    mice, only one dose per sex was used, and the number of mice (30) in
    the TCDD-exposed groups was considered less than optimal.

    7.7.2  Long-term animal studies with mixed compounds

         Toth et al. (1979) studied groups of 100 male and 100 female,
    10-week-old random-bred Swiss H/Riop mice that were given weekly oral
    doses of 2,4,5-trichlorophenoxyethanol (TCPE) at 67-70 mg/kg body
    weight, together with 0.112 mg TCDD/kg body weight or 0.007 mg TCDD/kg
    body weight in 0.5% carboxymethyl cellulose by gastric intubation for
    12 months. The incidences of liver tumours in males after 2 years were
    reported to be 48% and 58% in the two treated groups, compared with
    26-33% in the untreated male mice of the colony that survived up to 3
    years. Three additional groups of mice were given 7 g TCPE/kg body
    weight with 0.0007 g TCDD/kg body weight, 0.7 g TCPE/kg body weight
    with 0.00007 g TCDD/kg body weight, or 7 g TCPE/kg body weight with
    0.7 g TCDD/kg body weight. There was no increased incidence of liver
    tumours in any of the treatment groups.

         A 1:2 mixture of 1,2,3,6,7,8- and
    1,2,3,7,8,9-hexa-chlorodibenzo-p-dioxins (HxCDDs) has been tested for
    carcinogenicity by dermal application to mice and by gavage in rats
    and mice (NIH, 1980a,). The following impurities were detected in the
    mixture: pentaCDD 0.04%, TCDD 0.09%, triCDD 0.004%, and bromopentaCDD
    < 0.004%. The specific isomers of these impurities were not
    identified. The doses used and duration of the gavage studies (NIH,
    1980b) are given in Table 59. In both species and either sex, only
    tumours of the liver occurred at a significantly greater incidence
    than controls. In male rats and male and female mice, the liver tumour
    incidence was significantly increased over control values only in the
    high dose groups (5 g/kg per week), while in female rats the
    incidence was significantly greater at both medium- and high-dose
    levels (2.5-5 g/kg per week). In the dermal study, no
    treatment-related tumours were recorded in either the carcinogenicity
    bioassay or the tumour promotion assay using DMBA as an initiator
    (NIH, 1980a). It was concluded that the mixture of hexaCDDs tested was
    carcinogenic to rats and mice following administration by gavage.
    However, there was no tumorigenic activity when hexaCDD was applied to
    mouse skin.

         When added to the diet in concentrations up to 10 000 g/kg,
    2,7-dichlorodibenzo-p-dioxin and dibenzo-p-dioxin were found to be
    non-carcinogenic in chronic feeding studies in mice and rats of either
    sex (NCI, 1977; NCI, 1979).

    7.7.3  Short-term and interaction studies

         Poland & Glover (1979) estimated the maximum covalent binding of
    TCDD in vivo to rat liver protein, ribosomal RNA (rRNA), and DNA
    after 3H-TCDD (39 Ci/mmol) was administered to immature male and
    female Sprague Dawley rats (105-135 g) as a single ip injection of 7.5
    g/kg. The rats were killed 12 h, 24 h, 48 h, or 7 days after dosing
    with TCDD. The level of radioactivity in the liver varied from 18 to
    64% of the administered dose, and only a small fraction was associated
    with the purified macromolecular fractions. The radioactivity
    associated with rRNA and DNA was very low and essentially all the
    unextracted radioactivity was associated with protein (0.03 to 0.1% of
    the total radioactivity in the liver). The maximum amount of 3H-TCDD
    that could have been covalently bound to DNA was estimated as 1.8 x
    10-17 mol TCDD per mg DNA, or 6.2 nmol TCDD per mol DNA nucleotide,
    which means binding of about 1 molecule TCDD to the DNA in 35 cells.
    Phenobarbital treatment, or prior administration of TCDD did not
    significantly alter the amount of unextractable 3H-TCDD associated
    with any macromolecular fraction. Similarly, there were no differences
    in the levels of 3H-TCDD associated with protein, rRNA, or DNA in
    male or female rats pretreated with TCDD.

         TCDD was found to be a carcinogen in chronic feeding studies in
    rats and mice. Most carcinogens bind covalently, either directly or
    after a conversion to electrophilic intermediates, to protein, rRNA,
    and DNA to the extent of 10-4 to 10-6 mol of carcinogen per mol of
    amino acid or nucleotide residue. The maximum binding of TCDD is 4-6
    orders of magnitude lower than that of most chemical carcinogens and
    is of questionable biological significance. The results obtained in
    the study of Poland & Glover (1979) thus indicate that it is unlikely
    that the mechanism of TCDD-induced carcinogenesis would include the
    covalent binding of TCDD.

         In female Charles River CD-1 mice, TCDD was found to be a weak
    initiator when given alone in a single dose of 2 g/ mouse by dermal
    application (Di Giovanni et al., 1977). In these studies
    12-O-tetradecanoylphorbol-13-acetate (TPA) was used as a promoter.
    When TCDD and 7,12-dimethylbenzanthracene were given together, a
    slight additive effect was found. As mentioned earlier, in a study on
    a hexaCDD mixture, no treatment-related tumours were found in a tumour
    promotion test on mice using DMBA as an initiator (NIH, 1980a).

         The possible role of TCDD as a promoter in
    diethylnitrosamine-induced hepatocarcinogenesis was studied by Pitot
    et al. (1980) in female Charles River rats (200-250 g). A single oral
    dose (10 mg/kg)  of diethylnitrosamine (DEN) was given 24 h after a
    70% hepatectomy, and treatment with TCDD (0.14 or 1.4 mg/kg sc once
    every 2 weeks for 7 months) was started one week after the
    hepatectomy. The promoting effect of TCDD in this 2-stage model of
    liver cancer was also compared with the effect of a known promoting
    agent, phenobarbital (0.05% in the diet for 7 months). Enzyme-altered

    foci, which are thought to be precursors of hepatocellular carcinomas,
    were greatly increased in number, total volume, and phenotypic
    heterogeneity by the administration of TCDD. A significant incidence
    of hepatocellular carcinomas (5 out of 7) was observed in the
    DEN-treated rats that were given the high dose of TCDD, but no
    carcinomas were seen in the rats treated with DEN only (0 out of 4).
    The results indicated that TCDD was a potent promoting agent for
    hepatocarcinogenesis, and the authors suggested that all the tumours
    associated with the chronic administration of TCDD arise from its
    promoting activity of cells previously initiated by exposure to
    carcinogens in the environment.

         Studies utilizing a two-stage system of mouse skin tumorigenesis
    (Berry et al., 1979), which allows separate evaluation of the
    initiation and promotion phases of carcinogenesis, have demonstrated
    that TCDD does not promote the development of skin tumours at a dose
    of 0.1 g given twice weekly, whereas in the animals pretreated with
    1.0 g TCDD for 1, 3, or 5 days prior to initiation with DMBA, TCDD
    was shown to act as a potent inhibitor of PAH-induced skin tumour
    initiation. Almost complete inhibition (96%) was achieved with a
    single non-toxic topical dose of 0.1 g, and 3 days pretreatment with
    0.01 g TCDD gave over 80% inhibition. The authors suggested that this
    potent anticarcinogenic effect of TCDD may be related to its ability
    to induce epidermal enzyme pathways involved in detoxifying PAH
    carcinogens in the skin. According to Kimbrough (1979), TCDD and other
    compounds of this type, which are potent enzyme inducers, may prevent
    or enhance the tumour-inducing ability of other chemicals by enhancing
    the metabolism of these xenobiotics.

         Poland et al. (1982) studied the promoting effects of TCDD in the
    mouse skin two-stage tumorigenesis model. The effects of TCDD and TPA
    were compared in DMBA-initiated HRS/S mice that were either
    heterozygous or homozygous for the recessive "hairless" trait. TCDD
    was found to have a tumour-promoting effect only in the homozygous
    mice. The data suggested to the authors that TCDD might act as a
    promoter by a mechanism different from that of TPA.

         The interaction of TCDD with 3-methylcholanthrene (3-MC) was
    studied by Kouri et al. (1978), who found that TCDD was a
    co-carcinogen with 3-MC when administered by subcutaneous injection.
    Both sexes of two inbred strains of mice (C57Bl/6C and DBA/2),
    responsive and non-responsive to the induction of AHH by 3-MC,
    respectively, were used. TCDD at a concentration of 1 or 100 g/kg
    body weight was administered as a single dose alone or in combination
    with 3-MC (150 g/kg). The duration of the study was 36 weeks. The
    number of animals in each group at the start of the experiment was not
    stated, but seems to have been between 30 and 100. No subcutaneous
    tumours were observed in controls or in mice treated with TCDD alone.

    In responsive mice no enhancement occurred, while in non-responsive
    mice the simultaneous administration of TCDD and 3-MC enhanced the
    carcinogenic response of TCDD at 100 mg/kg. At 1 mg TCDD/kg, a
    reduction in latency time to tumour was noted.

         An anticarcinogenic effect of TCDD has been reported by Cohen et
    al. (1979) and by Di Giovanni et al. (1979a). When TCDD was topically
    applied to Sencar or CD-1 mice 72 h prior to the administration of
    either DMBA (10 nmol) or benzo(a)-pyrene (BP) (100 nmol), it markedly
    decreased the skin tumour initiation by both DMBA and BP. This
    inhibition of tumorigenesis correlated with the decreased in vivo
    binding of DMBA to DNA after TCDD administration, but not with the
    total binding of BP to DNA. However, the
    hydrocarbon-deoxyribonucleoside adducts from the DNA of
    TCDD-pretreated mice showed a striking absence of
    BP-7,8-dihyrodiol-9,10-epoxide adduct bound to guanine. It is
    suggested, accordingly, that the formation of this adduct may be a
    critical step in BP-induced skin carcinogenesis in mice. In further
    studies of the tumour-inhibitory effect of TCDD (Di Giovanni et al.
    1980), it was demonstrated that exposure of CD-1 mice to TCDD 3 days
    before initiation with BP or 3-MC resulted in a decreased tumour
    yield, compared to acetone-pretreated animals, while treatment with
    TCDD 5 min before and 1 day after initiation failed to affect the
    tumour yield. However, when TCDD was administered 3 days or 5 min
    before or 1 day after initiation with BP-diol epoxide, there was a
    decreased tumour yield in all cases. The authors concluded that the
    ability of TCDD to inhibit tumour yield when administered after the
    BP-diol epoxide, indicated the possible existence of more than one
    mechanism involved in the anticarcinogenic effect of TCDD.

    7.8  Mechanisms of Action

         The toxicity of TCDD apparently depends on the fact that the four
    lateral positions of the molecule are occupied by chlorine (see
    section 7.8.1) Toxicity decreases with decreasing lateral substitution
    and increasing total chlorine substitution. As has been outlined in
    sections 7.1-7.7, TCDD toxicity involves many different types of
    symptoms and these symptoms vary from species to species and from
    tissue to tissue, both quantitatively and qualitatively. Furthermore,
    age- and sex-related differences in sensitivity to TCDD have been
    reported. Characteristic for TCDD toxicity is also the delay in
    expression of toxicity, from 2 weeks to 2 months, seen in all species.
    It has been suggested that the initial event in TCDD-induced toxicity
    is the binding of TCDD to the so-called Ah receptor. This complex,
    whether of cytosolic or nuclear origin, exerts its action in the
    nucleus by triggering a pleiotropic response including the induction
    of mixed function oxidases. Present knowledge, however, rules out
    enzyme induction per se as being the cause of toxicity and death
    (see section 7.8.1). Although the toxicokinetics of TCDD vary between
    species, these differences are not sufficient to explain the
    variabilities in sensitivity to TCDD toxicity (see section 7.8.2).

    Available data indicate an involvement of TCDD in processes regulating
    cellular differentiation and/or division. Alterations in the
    regulation of such processes, which are not equally active in all
    cells throughout the organism, would be expected to result in effects
    that vary among tissues as well as among species (see section 7.8.3).

    7.8.1  Receptor-mediated effects

         The binding of TCDD to the Ah receptor has been postulated to be
    the necessary first step in the induction of cytochrome P-450
    synthesis and of related enzyme activities, as well as in the
    mechanism of toxicity (Poland et al., 1976; Okey et al., 1979; Poland
    & Glover, 1979; Poland & Knutson, 1982). So far, no conclusive data
    exist for the direct involvement of the Ah receptor in the
    TCDD-induced toxicity.

         Knowledge of the mechanism involved in the Ah locus
    enzyme-induction response has grown rapidly since the initial indings
    that binding of TCDD to the receptor resulted in increased levels of
    cytochrome P-450 mRNA in genetic variants of mice (Tukey et al., 1982)
    and mouse hepatoma cell lines (Israel & Whitlock, 1983). These
    findings have been confirmed and further expanded in studies using
    mice genetics and recombinant DNA techniques (Tukey et al., 1982;
    Miller et al., 1983; Gonzales et al., 1984; Israel & Whitlock, 1984;
    Jones et al., 1984, 1985, 1986; Okino et al., 1985; Tuteja et al.,
    1985; Kimura et al., 1986), thus providing more data to the
    understanding of the mechanism for the Ah locus enzyme induction.

         In early experiments (Poland et al., 1976; Carlstedt-Duke, 1979;
    Okey et al., 1979), Ah receptors appeared to be localized in the
    cytosol when in its unoccupied state and was translocated into nuclei
    only when occupied by a ligand (Greenlee & Poland, 1979; Okey et al,
    1980; Poellinger et al., 1982; Gasiewicz & Rucci, 1984). However,
    Whitlock & Galeazzi (1984) concluded that unoccupied Ah receptor in
    the intact cell was primarily located in the nucleus and that apparent
    cytosolic Ah receptor was a redistribution artifact. Following the
    distribution of Ah receptor and three cytosolic marker enzymes between
    the nuclear and cytosolic fractions during fractionation (Denison et
    al., 1986a,c), it was again concluded that unoccupied Ah receptor is
    primarily cytosolic or that this receptor protein is in equilibrium
    between the cytoplasm and nucleus.

         However, it is generally agreed that the ultimate biological
    regulation by the Ah receptor is due to specific interaction of
    ligand-receptor complexes with chromatin sites (Greenlee & Poland,
    1979; Okey et al., 1979, 1980; Mason and Okey, 1982; Poellinger et
    al., 1982; Poland and Knutson, 1982; Tukey et al., 1982a; Gonzales et
    al., 1984; Israel & Whitlock, 1984).

        Table 60. Physicochemical data for the hepatic Ah receptor in Sprague Dawley rats

                                                                               
    Physicochemical data       Denison et al. (1986a)  Poellinger et al. (1983)
                                                                               

    Stokes radius (nm)              5.2  0.2              6.1  0.2

    Sedimentation coefficient (S)   5.6  0.6              4.4

    Relative molecular mass         121 000                111 000

                                                                               
    
         Several investigators have estimated the molecular size and other
    physicochemical properties for the cytosolic hepatic Ah receptor (Okey
    et al., 1979, 1980, 1982; Tukey et al., 1982b; Poellinger et al.,
    1983; Gasiewicz et al., 1983a,b; Denison et al., 1986a; Hannah et al.,
    1986). To obtain reliable results in the isolation and
    characterization of the Ah receptor, it is necessary to use perfused
    liver, in order to reduce the contribution from blood proteins, and to
    use a radioactive ligand of high purity and specific activity quality
    (Poellinger et al., 1983; Denison et al., 1986a). Further more, the
    ionic strength of the medium during isolation has a marked effect upon
    the apparent molecular weight of the receptor (Denison et al., 1986a).
    The physicochemical data for the Ah receptor presented in Table 60
    were obtained from two studies (Denison et al., 1986a,c; Poellinger et
    al., 1983) in which the receptor was isolated from perfused liver of
    Sprague Dawley rats under conditions of high ionic strength.

         The receptor protein has been found also in extrahepatic tissues
    (Carlstedt-Duke, 1979; Carlstedt-Duke et al., 1979, 1981; Johansson et
    al., 1982; Mason & Okey, 1982; Gasiewicz &  Rucci,  1984; Gasiewicz et
    al., 1984; Furuhashi et al., 1986;  Kurl et al., 1985; Sderkvist et
    al., 1986). Different mammalian species possess Ah receptors with
    similar, though not identical, properties (Gasiewicz & Rucci, 1984;
    Denison et al., 1986a,b; Kurl et al., 1985). Jaiswal et al. (1985a,b)
    have shown species differences in the TCDD-inducible P-450 gene
    subfamily. Humans appear to only have the P1-450. The function of
    human P1-450 may be equivalent to a combination of  P1-450 and P3-450
    in the mouse. A complete lack of measurable  cytosolic and almost
    total absence of inducer-receptor complexes in the nucleus of human
    MCF-1 cells (cells derived from an adenocarcinoma of the breast) were
    reported. This absence was out of proportion to the ability of TCDD to
    induce  AHH and acetamide-4-hydroxylase activities in these cells.
    Further studies in different cell lines are thus needed to
    characterize the level of receptor in humans. The only non-mammalian

    species demonstrated to have significant Ah receptor  is the chick
    embryo (Denison et al., 1986b). However, no detectable level of the
    receptor was found in 2-week-old White  Leghorn chickens (Sawyer et
    al., 1986). Based on certain similarities in the biochemical behaviour
    between the Ah receptor  and steroid hormone receptors, it has been
    proposed that there  is a natural ligand for the Ah receptor (Neal et
    al., 1979; Poland et al., 1976). So far such a ligand has not been
    identified, either among steroid hormones (Poland et al., 1976;
    Carlstedt-Duke et al., 1979; Romkes et al., 1987) or among certain
    dietary factors (Johansson et al., 1982), although lumichrome, a
    metabolite of riboflavin, was suggested as an endogenous ligand for
    the receptor (Kurl & Villee, 1985). TCDD  does not bind to the
    glucocorticoid, estrogen or progesterone  receptors (Neal et al.,
    1979; Romkes et al., 1987). Monoclonal  anti-glucocorticoid
    receptor-IgG antibodies did not react with  the TCDD receptor
    (Poellinger et al., 1983) and the hydrophobic properties of the Ah
    receptor were more pronounced than  those of the steroid hormone
    receptors (Poellinger & Gullberg,  1985).

         Convincing data for the importance of the receptor in
    TCDD-induced toxicity could be based on structure activity
    relationships, i.e., that the binding affinities of TCDD and other
    PCDDs or PCDFs to the receptor correlate with their biological
    potencies. The binding affinities of PCDDs and PCDFs have been
    demonstrated to correlate with their biological potencies,
    particularly the induction of enzyme activities as well as the
    production of acute toxic effects (Tables 56 and 61) (Poland & Kende,
    1976; Poland et al., 1976; Knutson & Poland, 1982).

         Furthermore, the structure-activity relationships observed for
    enzyme induction, thymic atrophy, body weight loss, and LD50 values
    were comparable to the structure-activity relationships observed for
    receptor binding (Tables 56, 61) (Bandiera et al., 1984a,b; Mason et
    al., 1985, 1986; Sawyer & Safe, 1985; Safe et al., 1986). Interactive
    studies, i.e., studies where PCDD and PCDF congeners have been given
    both separately and as mixtures, have also been used to investigate
    the role of the Ah receptor in the mechanism of action of TCDD.
    Depending on the mechanism of action the biological responses may be
    synergistic potentiated, additive, unaffected, or antagonistic. Such
    studies have been performed for enzyme induction (Sawyer & Safe, 1985;
    Keys et al., 1986; Ahlborg et al., 1987), vitamin A reduction
    (Hakansson et al., 1987), teratogenicity (Birnbaum et al., 1985; Weber
    et al., 1985), thymic atrophy (Bannister & Safe, 1987), and immune
    suppression (Rizzardini et al., 1983). So far this kind of data is
    scattered and difficult to interpret.



    
    Table 61.  Structure-activity relationships for some PCDFs
                                                                                                                            

                   In vitro EC50 values (mol/litre)a,b            In vivo ED50 values  (mol/kg)              LD50 c
                                                                                                                            

    PCDF           Receptor            AHH                 EROD           AHH            weight    Thymic    Guinea-pig
    congener       binding                                                               loss      atrophy   g/kg body weight
                                                                                                                            

    Dibenzofuran   < 10-3            ND                  ND
    2-             2.8  x 10-4       ND                  ND
    3-             4.2  x 10-5       ND                  ND
    4-             < 10-3            1.0  x 10-5       1.71 x 10-5
    2,3-           4.72 x 10-6       2.19 x 10-6       4.84 x 10-6
    2,6-           2.46 x 10-4       6.17 x 10-5       6.31 x 10-5
    2,8-           2.57 x 10-4       3.95 x 10-5       4.0  x 10-5
    1,3,6-         4.40 x 10-6       2.53 x 10-6       3.37 x 10-6
    1,3,8-         8.50 x 10-5       1.94 x 10-5       3.02 x 10-5
    2,3,4-         1.9  x 10-5       1.51 x 10-7       2.48 x 10-7
    2,3,8-         1.0  x 10-6       2.49 x 10-6       1.56 x 10-6
    2,6,7-         4.5  x 10-7       2.80 x 10-6       3.13 x 10-6
    2,3,4,6-       3.5  x 10-7       1.32 x 10-6       1.13 x 10-6
    2,3,4,7-       2.51 x 10-8       1.79 x 10-8       1.48 x 10-8       46             34      7.8
    2,3,4,8-       2.0  x 10-7       4.14 x 10-8       3.76 x 10-8       ND            130     > 150
    2,3,6,8-       2.2  x 10-7       1.04 x 10-6       7.79 x 10-7
    2,3,7,8-       4.1  x 10-8       3.91 x 10-9       2.02 x 10-9     0.65           3.2       3.6            5-10
    1,2,3,6-       3.54 x 10-7       > 10-4            > 10-4         > 160          > 250     > 250
    1,2,3,7-       1.12 x 10-7       2.7  x 10-5       6.3  x 10-5      110             87      110
    1,2,4,8-       > 10-5            1.20 x 10-5       9.26 x 10-5
    1,2,4,6,7-     6.77 x 10-8       3.25 x 10-7       3.48 x 10-7
    1,2,4,7,9-     2.0  x 10-5       3.77 x 10-8       3.84 x 10-8
    1,2,3,4,8-     1.2  x 10-7       2.09 x 10-7       1.63 x 10-7
    1,2,3,7,8-     7.45 x 10-8       2.54 x 10-9       3.06 x 10-9     1.5            2.6       1.8
    1,2,3,7,9-     3.98 x 10-7       8.6  x 10-8       8.6  x 10-8       15             49       23
    1,2,4,6,8-     3.09 x 10-6       1.0  x 10-5       1.2  x 10-5    > 150           7150     > 150

                                                                                                                            

    Table 61 (contd).
                                                                                                                            
                   In vitro EC50 values (mol/litre)a,b            In vivo ED50 values  (mol/kg)              LD50 c
                                                                                                                            

    PCDF           Receptor            AHH                 EROD           AHH            weight    Thymic    Guinea-pig
    congener       binding                                                               loss      atrophy   mg/kg body weight
                                                                                                                            
    
    1,2,4,7,8-     1.3  x 10-6       1.06 x 10-7       1.48 x 10-7       7.8              49       46
    1,3,4,7,8-     2.0  x 10-7       1.60 x 10-9       1.40 x 10-9       3.5              26        0.70
    2,3,4,7,8-     1.5  x 10-8       2.56 x 10-10      1.34 x 10-10      0.037             1.0      0.26
    2,3,4,7,9-     2.0  x 10-7       7.9  x 10-9       5.8  x 10-9       7.0              22        5.5
    1,2,3,4,7,8-   2.3  x 10-7       3.56 x 10-10      3.79 x 10-10      0.29              1.3      0.56
    1,2,3,6,7,8-   2.7  x 10-7       1.47 x 10-9       1.24 x 10-9       0.35              3.2      0.90
    1,2,4,6,7,8-   8.3  x 10-6       4.24 x 10-8       2.93 x 10-8
    2,3,4,6,7,8-   4.7  x 10-8       6.87 x 10-10      5.75 x 10-10      0.27              2.8      0.90
                                                                                                                            

    a     Estimated concentration needed to displace 50% of 3H-TCDD bound to liver cytosol receptor from
          Wistar rats and to produce 50% maximum enzyme induction in the rat hepatoma H-4-IIE cell line (Bandiera
          et al., 1984b).
    b     Studies in immature male Wistar rats (Mason et al., 1985).
    c     Moore et al. (1979).
    


         Polymorphism in the Ah locus, which is suggested to be structural
    gene for the cytosolic receptor, seems to determine the sensitivity of
    genetically different strains of mice to TCDD and congeners. Ah
    responsive strains of mice, e.g., C57Bl/6, are characterized by (a)
    high hepatic levels of the TCDD receptor protein, (b) highly elevated
    levels of hepatic cytochrome P-448 and associated enzyme activities in
    response to treatment with 3-MC, and (c) sensitivity to the ulcerative
    action of DMBA on the skin. Ah-non-responsive mice, e.g., DBA/2, lack
    these attributes (Nebert et al., 1975). Based on these findings
    several genetic studies have been performed to elucidate the role of
    the receptor in TCDD toxicity. Contrary to 3-MC, TCDD induces AHH
    activity and several toxic effects both in Ah-responsive and
    Ah-non-responsive strains of mice. However, the dose required to
    produce the effect in an Ah-non-responsive strain is approximately
    10-fold greater than that needed for a responsive strain, thus
    demonstrating that the Ah-non-responsive strain also contains the
    TCDD-receptor but that this receptor is defective (Okey & Vella,
    1982).

         Crosses and backcrosses of C57BL/6 and DBA/2 mice have shown that
    sensitivity to TCDD-induced thymic atrophy immune system disturbances
    (section 7.4.5) and teratogenic effects (section 7.5.2) segregate with
    the Ah locus. Furthermore, data from studies of DBA/2 mice given
    either single or multiple doses of TCDD (Jones & Sweeney, 1980; Smith
    et al., 1981) suggest that the LD50 in this strain of mice is at
    least 5-fold greater than the values recorded for the C57Bl/6 and
    C57Bl/10 strains (Vos et al., 1974; Jones & Greig, 1975; Smith et al.,
    1981). TCDD-induced hepatic porphyria has also been shown to segregate
    with the Ah locus in mice (Jones & Sweeney, 1980). However, Greig et
    al. (1984) found that additional genetic loci must be involved in this
    lesion. The correlative differences between the C57Bl/6 and DBA/2
    strains of mice, in terms of altered specific binding of TCDD and
    sensitivity to this compound, may be unique and may not be applicable
    to other species (Gasiewicz & Rucci, 1984).

         Less convincing data for the model of receptor-mediated toxicity
    of TCDD arise from studies of toxicity, receptor levels, and/or enzyme
    induction of TCDD in various species, tissues, and cell cultures.
    Despite enormous variability in recorded LD50 values for guinea-pig,
    rat, mouse, rabbit, and hamster (Table 47), the amounts and physical
    properties of the hepatic as well as the extrahepatic receptors, do
    not vary extensively in these species (Poland & Knutson, 1982;
    Gasiewicz & Rucci, 1984). Furthermore, although recorded LD50 values
    for TCDD vary more than 100 times in chick embryos, C3H/HeN mice, and
    Sprague Dawley rats, the ED50 doses for AHH induction in these
    species are comparable (Poland & Glover, 1974b). In the guinea-pig,
    the most TCDD-susceptible species, enzyme induction is several times
    lower even at lethal doses. A number of cell types, including primary

    cultures and established and transformed cell lines from several
    species and tissues, are inducible for AHH activity, indicating the
    presence of the receptor, yet toxicity is not expressed in these
    systems (Knutson & Poland, 1980a).

         Available data thus suggest that the receptor for TCDD may be a
    prerequisite, but is not sufficient in itself for the expression of
    TCDD toxicity.

    7.8.2  Toxicokinetics

         The interspecies variation in sensitivity to TCDD may be
    attributable, at least in part, to different rates at which various
    species distribute, metabolize, and excrete the compound. Table 62
    summarizes some of the data on the elimination, toxicity, and
    metabolism of TCDD in different species, some previously discussed in
    section 6. Distribution data (section 6.1) has been obtained mainly
    from animals exposed to toxic doses of TCDD. Interspecies comparisons
    based on these data are difficult to perform, since studies in
    different species have been performed with non-comparable relative
    toxic doses and collection of data has occurred at variable time
    points. However, the available data suggest that tissue levels alone
    cannot explain the interspecies differences in pathology and acute
    toxicity of TCDD. For example, the molar concentration of TCDD in the
    hamster may be orders of magnitude greater than in the rat and mice,
    without development of hepatotoxicity, yet definite liver damage
    occurs both in rats and mice.

         Based on the findings that the toxicity of TCDD is lower in rats
    after the stimulation of hepatic mixed function oxidases (Beatty et
    al., 1978), and that the metabolites of TCDD are less toxic than the
    parent compound (Weber et al., 1982a; Mason & Safe, 1986) (see section
    7.1.5), the metabolism of TCDD has been considered as a detoxification
    mechanism.

         Following the administration of TCDD, most of it appears to be
    eliminated through a first-order process in most species (see section
    7.1). In all species investigated TCDD is largely eliminated in the
    faeces (Table 48). Only in hamsters and certain strains of mice is
    urinary elimination a major route of excretion (Table 41). Available
    data demonstrate that TCDD is converted to more polar metabolites
    prior to elimination in the urine and bile (Table 62). Unchanged TCDD
    does not appear in the bile or urine of any species, but it is the
    major excretory product in the faeces of mice, guinea-pigs, and
    hamsters (Table 62). As the urine and bile appear to be free of
    unmetabolized TCDD, the existence of unchanged TCDD in faeces
    indicates that a significant amount of unchanged TCDD may be excreted
    into the intestinal lumen by some route other than bile.



    
    Table 62. Rates of elimination, toxicity, and metabolic transformation of TCDD in different species
                                                                                                                            
    Species/strain      Elimination    LD50 value                  TCDD-derived radioactivity occurring in:
                        half-life      (g/kg
                        (days)         body weight)        Urine          Bile                Faeces         Tissues
                                                                                                                            

    Ratsa,b,c,d

    Sprague Dawley         17-31       25-60               5 polar        4-8 polar                          unchanged TCDD
                                                           metabolites    metabolites;
                                                                          little, if any,
                                                                          unchanged TCDD.

    Micee,f,g

    C57BL/6, DBA/2,        10-24       114-2570            4-7 polar      4-6 polar           3-4 polar      unchanged TCDD
    B6D2F1                                                 metabolites    metabolites         metabolites,
                                                                                              unchanged TCDD.

    Guinea-pigh

    Hartley                22-94       0.6-2.5             4 polar        5 polar             mainly         unchanged TCDD,
                                                           metabolites    metabolites         unchanged      polar metabolites
                                                                                              TCDD
    Hamstersa,i

    Golden Syrian          12-15       1157-5051           4 polar        5-6 polar           metabolites,   unchanged TCDD
                                                           metabolites    metabolites         unchanged
                                                                                              TCDD

    Dogsj

    Beagle              not reported   not reported                       metabolites
                                                                                                                            

    a Gasiewicz et al., 1983a.  b Poiger & Schlatter, 1979.  c Ramsey et al., 1982.      d Rose et al., 1976.
    e Gasiewicz et al., 1983b.  f Koshakji et al., 1984.     g Vinopal & Cassida, 1973.  h Olson, 1986.
    i Olson et al., 1980a.      j Poiger et al., 1982.
    

         The metabolic profiles of TCDD in excreta differs between
    species, but generally urinary metabolites are more polar than biliary
    or faecal metabolites. Furthermore, several metabolites, both in urine
    and bile, are glucuronide conjugates.

         The apparent absence of TCDD metabolites in the tissues of all
    species, except for the guinea-pig (Table 62), suggests that, once
    formed, the metabolites of TCDD are readily excreted. Another factor,
    besides metabolism, that may influence the total rate of elimination
    of TCDD is the amount of adipose tissue stores, which may vary between
    species.

         At present there is no clear relationship between the ability of
    a given species to excrete TCDD and/or its metabolites and the acute
    toxicity of TCDD in that species. However, the somewhat greater rate
    of elimination of TCDD in the hamster and the lower rate of
    elimination in the guinea-pig (Table 62) may contribute to their
    relative resistance and sensitivity, respectively, to the acute toxic
    effects of TCDD.

         The rate of metabolism and excretion of PCDDs and PCDFs varies
    with molecular structure. In most species PCDFs are much more readily
    eliminated than their PCDD counterparts. Less halogenated congeners
    are usually metabolised and excreted more rapidly than the more
    halogenated ones, especially the 2,3,7,8-substituted congeners (see
    sections 6 and 9).

    7.8.3  Impairment of normal cellular regulatory systems

         When considered together, the diverse pattern of toxic effects,
    the species and tissue-specific responses, and the time-course for
    effects, as well as the non-toxic action of TCDD on most cell cultures
    in vitro, seem to indicate that TCDD-induced toxicity occurs as a
    result of an impairment of a normal cellular regulatory system. Such
    a system might be present in all cells throughout the organism, though
    the activity may vary with cell type, tissue, age, sex, and strain,
    and/or species.

    7.8.3.1  Endocrine imbalance

         In many aspects, TCDD toxicity mimics endocrine imbalance,
    although no evidence exists to indicate a direct involvement of
    hormones in the toxic action of TCDD (section 7.4.9).

    7.8.3.2  Body weight regulation

         The most reliable and consistent symptom of TCDD toxicity among
    all experimental animals is weight loss. The cause of the body weight
    loss seems to be reduced food intake, apparently occurring secondarily
    to a physiological adjustment that reduces the body weight to a

    maintenance level lower than normal. The physiological trigger for
    controlling this body weight set-point might be a target for TCDD
    action (section 7.4.1).

    7.8.3.3  Plasma membrane function

         The changes in the surface characteristics of the plasma
    membranes induced by TCDD in vivo (7.4.2.2) resemble changes
    occurring in precancerous and transformed cells (Pitot & Sirica,
    1980). Such changes, including reduction of gap junctions and surface
    glyco-proteins, would be expected to curtail cell-cell communication
    and to reduce intercellular recognition and attachment events
    implicated in the process of tumour promotion.

         It has been shown by Pitot et al. (1980) that TCDD promotes
    diethylnitrosamine (DEN)-induced hepatocarcinoma in rats. In this
    study, canalicular ATPase was used as a marker in detecting
    enzyme-altered foci, whose number increased when TCDD was given to
    DEN-treated partially hepatectomized rats. The foci exhibited
    decreased ATPase activity in agreement with previous observations that
    TCDD in vivo reduces the ATPase level in canaliculi-rich plasma
    membranes.

         TCDD, unlike other well known promoters, requires a prolonged
    treatment period in vivo to exert its effect. The lack of effect
    of TCDD in vitro would imply that the promoter activity is
    mediated through some in vivo process and not by its direct
    interaction with plasma membranes.

    7.8.3.4  Impaired vitamin A storage

         The histomorphological appearance of chloracne, the most
    characteristic and prominent sign of TCDD-induced toxicity in humans,
    resembles in some respects effects seen in the skin of patients
    suffering from vitamin A deficiency (Kimbrough, 1974). Many of the
    effects of TCDD poisoning observed in animal studies, including
    failure of normal growth, keratosis, epithelial lesions,
    immunosuppression, and reproductive and teratological effects, are
    similar to the effects of dietary vitamin A deficiency (Thunberg et
    al., 1980). The most intriguing similarities between symptoms due to
    vitamin A deficiency and TCDD toxicity concern effects on epithelial
    tissues, particularly the process of keratinization. TCDD induces
    terminal differentiation of epithelial tissues both in vivo and
    in vitro. However, lack of epithelial degeneration (programmed
    cell death) of the medial epithelial cells of palatal shelves has been
    reported in mice exposed to TCDD in utero (Pratt et al., 1984).

         Vitamin A is essential for normal differentiation. It diminishes
    the expression of differentiation in stratified squamous epithelia and
    accentuates the expression of differentiation in secretory epithelia.
    Vitamin A deficiency can convert secretory epithelia to squamous
    epithelia, while excess of the vitamin can convert stratified squamous
    epithelia to secretory epithelia (Wolf, 1980). With the use of
    cultured human keratinocytes, it has been demonstrated that vitamin A
    at the cellular level affects cell motility, cell-cell interaction,
    and epithelial morphogenesis. At the molecular level, vitamin A
    determines, by controlling the level of the corresponding mRNA, the
    nature of keratins synthesized (Fuchs & Green, 1981). Keratins
    constitute a cytoskeleton in epithelial cells, and the keratin pattern
    may be used as a marker for epithelial differentiation (Sun et al.,
    1979, 1983a,b). Removal of vitamin A from the medium of cultivated
    human keratinocytes of various origin led to increased synthesis of
    large keratins and reduced synthesis of lower molecular weight
    keratins (Fuchs & Green, 1981). This pattern was reversed by the
    addition of vitamin A to the medium. Each tissue and cell type
    controlled its synthesis of keratins differently, depending on the
    vitamin A concentration in the medium. The ability of TCDD to impair
    vitamin A storage (section 7.4.10) may be responsible for some of the
    toxic effects produced by TCDD.

    7.8.4  Lipid peroxidation

         Based on indirect lines of evidence, Sweeney & Jones (1983)
    proposed that increased in vivo lipid peroxidation, resulting in
    the formation of free radicals, might be a mechanism of TCDD toxicity.
    Firstly, lipofuscin pigments, considered to be by-products of lipid
    peroxidation, accumulate in the heart muscle cells of TCDD-treated
    rats (Albro et al., 1978). Secondly, iron deficiency inhibits in
    vitro lipid peroxidation (Bus & Gibson, 1979; Sweeney et al., 1979)
    and has been demonstrated to reduce hepatic TCDD toxicity in vivo
    in rats (Sweeney et al., 1979). Thirdly, 0.25% butylated
    hydroxyanisole (BHA) in the diet has been shown to provide protection
    from TCDD-induced porphyria and lipid accumulation in mice. In
    contrast, 0.01% vitamin E, another antioxidant, in the diet had no
    protective effect (Hassan et al., 1985a,b). Stohs et al. (1983)
    demonstrated increased in vivo (conjugated diene method) and in
    vitro (microsomal malondialdehyde formation) hepatic lipid
    peroxidation in female Sprague Dawley rats given a total of 70 g/kg
    body weight in three daily oral doses of 10, 20, and 40 g TCDD/kg
    body weight, or a single oral dose of 80 g TCDD/kg body weight. Lipid
    peroxidation was determined at days 1, 6, and 11 after the last
    treatment. The maximal increase of lipid peroxidation in vivo was
    2-fold one day post-treatment, whereas the 5- to 6-fold increase in
    in vitro lipid peroxidation reached its maximum at 6 days

    post-treatment. The TCDD-induced in vitro lipid peroxidation could
    be inhibited by repeated treatment with BHA, glutathione, vitamin E,
    and vitamin A (Hassan et al., 1985a,b). Dietary selenium had no
    inhibitory effect on TCDD-induced lipid peroxidation (Hassan et al.,
    1985c).

         Robertson et al. (1985) found no evidence for TCDD-induced in
    vivo lipid peroxidation, as judged by levels of exhaled endogenous
    ethane, and metabolic clearance of both externally and internally
    applied exogenous ethane in male Sprague Dawley rats after a single ip
    dose of 60 g TCDD/kg body weight. Neither was there a correlation
    between TCDD-induced lipid peroxidation in vitro and sensitivity
    towards the lethal effect of TCDD in Sprague Dawley rats, Golden
    Syrian hamsters or guinea-pigs (Hassan et al., 1983). Hepatic
    microsomes from Sprague Dawley rats and Golden Syrian hamsters exposed
    to TCDD in vitro responded with increased lipid peroxidation only
    in the presence of Fe3+-ADP in the incubation mixture (Albro et al.,
    1986).

         From all these data on TCDD and lipid peroxidation, it was
    concluded by Albro et al. (1986) that it was premature to attempt to
    define a relationship between lipid peroxidation and TCDD-induced
    lethality.

    8.  EFFECT OF PCDDs ON HUMAN BEINGS - EPIDEMIOLOGICAL AND
    CASE STUDIES

    8.1  Occupational Studies - Historical Perspective

         The illness most frequently observed in workers engaged in the
    manufacture of trichlorophenol, 2,4,5-T, and related products is a
    skin disease called chloracne. This skin disease has also been called
    "Pernakrankheit" (perchlorinated naphthalene illness or halogen wax
    acne) and was described by Herxheimer (1899). In addition to the
    halogenated phenols, chloracne is caused by a number of chlorinated
    compounds such as the chlorinated biphenyls and chlorinated
    naphthalenes (Muller, 1937; Braun, 1955; Crow, 1970; Kimbrough, 1974).

         Although chloracne is well known to those engaged in the 
    treatment of occupational diseases, many outbreaks that have occurred
    over the years, particularly in the USA, have not been reported in the
    scientific literature. In the Federal Republic of Germany, chloracne
    is now considered an occupational disease for which compensation is
    mandatory (Braun, 1970).

         Herxheimer (1899) also described general toxic signs and symptoms
    in his patients, such as lack of appetite, weight loss, headache, and
    vertigo. After his original observations and publication, several
    other reports followed. The technique of obtaining chlorine gas
    consisted of an electrolytic procedure where a mixture of potassium,
    sodium, and magnesium chlorides was subjected to a current with a
    central carbon electrode where the chlorine was obtained and piped
    off. The workers who took care of the chlorine gas never developed
    chloracne thus refuting the original hypothesis by Herxheimer. By
    contrast, those who handled the electrodes and cleaned the reaction
    vessels were those afflicted with chloracne. Already at this time
    chlorinated phenolic compounds were considered as possible noxious
    agents (Fraenkel, 1902). This however could never be proven and even
    at present, when satisfactory analytical techniques are now available,
    no analysis of the so-called "tuffy tar" has been carried out.

         Another class of chlorinated organic compounds causing skin
    damage appeared during the First World War (1914-1918). At this time
    perchlorinated naphthalenes had come into use as insulation materials,
    e.g., in the radio and electronic industry. The first description of
    Pernakrankheit was that by Wauer (1918). The use of the unspecified
    technical mixture of chlorinated naphthalenes spread all over the
    world and caused numerous intoxications notably among workers in
    manufacture. The perna disease has been summarized by von Wedel et al.
    (1943) and described in particular detailed by Braun (1955). Apart
    from chloracne the systemic effects of the same compounds have been
    dealt with by Drinker et al. (1937) and Greenburg et al. (1939).

         Both in man and experimental animals, serious liver damage
    occurred after exposure to chlorinated naphthalenes, consisting of
    liver necrosis and toxic jaundice (acute yellow liver atrophy). Among
    several hundred cases of chloracne due to these compounds, Braun
    (1955) tabulated 24 deaths due to toxic jaundice and 14 recoveries. It
    should be pointed out that a fulminant liver disease with jaundice of
    this kind is an extremely rare condition. For comparison, it has never
    occurred after exposure to trichlorophenol (TCP) and TCDD as described
    below. Note should also be taken of the fact that not only were the
    perchlorinated naphthalenes an ill identified mixture of chemical
    species, but exposure frequently occurred at the same time to mixtures
    of chlorinated biphenyls, the latter now known to be contaminated with
    chlorinated dibenzofurans. The potentiation of toxicity by these
    mixtures and other chlorinated compounds were discussed by Drinker et
    al. (1937), Greenburg et al. (1939), von Wedel et al. (1943), and
    Risse-Sunderman (1959).

         Several accidental ingestions of chloracnegenic compounds have
    occurred. They are of particular importance in relation to discussions
    on whether chloracne is a systemic or local disease. The so-called
    Yusho disease is discussed in section 11.

         Herzberg (1947) described several cases of chloracne, in which
    other toxic signs and symptoms were seen, due to consumption of
    "chlorinated paraffin" used as a substitute for butter during cooking
    in postwar Berlin. Among general signs and symptoms observed were
    gastrointestinal disturbances with abdominal pain, headache, pain in
    joints, neuropathy, depression, and lack of appetite. The
    dermatological symptoms were erythema, exanthema, comedones, and
    retention cysts in sebaceous glands. It was noted as remarkable that
    the skin signs had a follicular predilection, as in seborrhoea (face,
    head, bosom, and back). The slow development of chloracne, and
    particularily the fact that the sebaceous glands were affected, led
    the author to conclude that it was a secretory disease
    (Ausscheidungstoxikose). With regard to the chloracnegenic component,
    it is unlikely that paraffin itself was active. Herzberg speculated
    that something else, possibly a pyrolysis product that arose during
    cooking, could have caused the disease.

         Accidents in chemical plants involved in the manufacture of
    chlorinated phenolic compounds are listed in Table 63. It should be
    stressed that all these intoxications are due to mixtures, e.g., TCP
    and TCDD and other compounds. Summaries of the industrial accidents
    are to be found in Holmstedt (1980) and only some of them will be
    dealt with here.



    
    Table 63. Summary of accidents in chemical plants involving the manufacture of chlorinated phenolic compounds
                                                                                                                                          
                                                                               Years from
                                            Cause of            Personnel      incident to
    Country and date         Producta       exposure            affected       last observation         References

                                                                                                                                         
    Germany 1910             CP             Explosion +           5            Same year                Teleky (1913), Wahle (1914),
                                            occupational                                                Dohmeier & Janson (1983)

    United States 1949       TCP            Explosion +         228            30                       Ashe & Suskind (1949, 1950),
                                            Occupational                                                Suskind et al. (1953),
                                                                                                        Suskind (1978), Huff et al.
                                                                                                        (1980), Zack & Suskind (1980)
                                                                                                        Zack & Gaffey (1983), Moses
                                                                                                        et al. (1984), Suskind &
                                                                                                        Hertzberg (1984)

    Federal Republic         TCP            Occupational         17            1                        Baader et al. (1951)
    of Germany 1949          (PCP)

    Federal Republic         TCP            Occupational         60                                     Bauer et al. (1961)
    of Germany 1952

    Federal Republic         TCP            Occupational         37                                     Hay (1977)
    of Germany 1952-1953

    Federal Republic         TCP            Explosion +          75            29                       Hoffman (1957), Goldmann (1972,
    of Germany 1953                         Occupational                                                1973), Huff et al. (1980), Thiess
                                                                                                        et al. (1982)

    France 1953              TCP            Explosion +          17            2                        Dugois & Colomb (1956, 1957),
                                            Occupational                                                Dugois et al. (1958)

    Federal Republic         TCP,           Occupational         31            24                       Schultz (1957a,b), Bauer et al.
    of Germany 1954          2,4,5-T                                                                    (1961), Kimmig & Schultz (1957a,b)
                                                                                                        Kleu & Gltz (1971), von Krause &
                                                                                                        Brassow (1978)
                                                                                                                                         

    Table 63 (contd - 2).
                                                                                                                                         
                                                                               Years from
                                            Cause of            Personnel      incident to
    Country and date         Producta       exposure            affected       last observation         References
                                                                                                                                         
    Federal Republic         TCP            Occupational         24            4                        Risse-Sundermann (1959)
    of Germany 1954          2,4,5-T

    United States 1956       2,4,5-T        Occupational         48            6                        Bleiberg et al. (1964), Poland
                             2,4,5-T                                                                    et al. (1971)

    United States 1956       TCP            Occupational        Many                                    Hay (1977)

    Italy 1959               TCP            Explosion +           5            2                        Hofman et al. (1962)
                                            Occupational

    United States 1959       TCP            Occupational                                                Hay (1977)

    United States 1960       TCP            Occupational        Many                                    Hay (1977)

    Netherlands 1963         TCP            Explosion           106            11                       Dalderup, (1974a,b), Berlin
                             2,4,5-T                                                                    et al. (1976), Huff et al. (1980)

    USSR 1964                2,4,5-T        Occupational        128                                     Telegina et al. (1970), IARC
                                                                                                        (1977)

    United States 1964       TCP            Occupational         61            6                        Vahrenholt (1977), Cook et al.
                                                                                                        (1980), Ott et al. (1980)

    Czechoslovakia 1964-1969 TCP            Occupational         80            6                        Jirsek et al. (1973, 1976),
                                                                                                        Pazderova et al. (1974, 1980,
                                                                                                        1981)

    United Kingdom 1968      TCP            Explosion            90            14                       May (1973, 1982), Huff et al.
                                                                                                        (1980)

    Japan 1970               2,4,5-T        Occupational         25            3                        Mivra et al. (1974)
                                                                                                                                         

    Table 63 (contd - 3).
                                                                                                                                        
                                                                               Years from
                                            Cause of            Personnel      incident to
    Country and date         Producta       exposure            affected       last observation         References
                                                                                                                            
    USSR 1972                TCP            Occupational          1            1                        Zelikov & Danilov
                                                                                                        (1974)

    Austria 1972-1973        2,4,5-T        Occupational         50                                     Forth (1977), Hay
                                                                                                        (1977)

    Federal Republic         2,4,5-T        Occupational          5                                     Forth (1977), Hay
    of Germany 1974                                                                                     (1977)

    Italy 1976               TCP            Explosion           193            8                        Reggiani (1977, 1978, 1983a),
                                                                                                        Vahrenholt (1977), Filippini
                                                                                                        et al. (1981), Abate et al.
                                                                                                        (1982), Ideo et al. (1982)
                                                                                                                             

    a     Products: TCP = trichlorophenols;  CP = chlorophenols;  PCP =  pentachlorophenol;
          2,4,5-T = 2,4,5-trichlorophenoxyacetic acid.
    


         The first reported intoxication with a mixture probably
    containing TCDD, although the chemical structure was not given,
    occurred in February 1910. Five people were said to have been
    contaminated after a reactor explosion and two of these were described
    in some detail in a dermatological thesis (Teleky, 1913; Wahle 1914).
    Wahle (1914), however, in his thesis emphazised that this intoxication
    was not due to any of the chlorinated naphthalene derivatives that
    were well known by then.

         An industrial poisoning was reported in 1949, due to the
    formation of TCDD in uncontrolled exothermic reactions occurring
    during the manufacture of TCP at a 2,4,5-T-producing factory in Nitro,
    West Virginia, USA. The temperature in one of the reactors containing
    tetrachlorobenzene, methanol, and sodium hydroxide increased, a relief
    valve opened, and the contents of the vessel were discharged into the
    interior of the building and over a wide area outside of the building.
    A total of 228 people were affected.

         Symptoms included chloracne, nausea, vomiting, headaches, severe
    muscular aches and pains, fatigue, emotional instability, and
    intolerance to cold. Laboratory findings showed an increase in total
    serum lipids and an initially prolonged prothrombin time. Among those
    affected were not only workmen, but also laboratory personnel, medical
    personnel, and even the Safety Director who visited the area of
    exposure. Several wives who had never visited the plant also developed
    acne, usually at the same time as their husbands working at the plant.
    A man from the nearby town who purchased a truck that was parked in
    the vicinity of the accident at the time it occurred, and his child,
    also developed chloracne. The disabling symptoms, which kept men from
    their jobs for as long as 2 years, were severe aches and pains and
    fatigability, the manifestations of peripheral neuropathy. Liver tests
    4 years later were normal, but mild cases of acne were common. TCDD
    was still an unknown chemical. The follow-up to this accident will be
    discussed in section 8.4.

         In 1953, at the Badische Anilin and Soda Fabrik, during the
    alkaline hydrolysis of 1,2,4,5-tetrachlorobenzene to
    2,4,5-trichlorophenol, the temperature and pressure in an autoclave
    increased rapidly and resulted in an exothermic reaction releasing a
    great deal of steam through a safety valve of the reaction vessel.
    This steam covered the walls, windows, doors, and machinery in the
    rooms of four floors, and finally precipitated in solid form on
    everything in these rooms. Forty-two workers involved in the clean-up
    operations developed chloracne, and even after the extensive clean-up
    operations occasional workers still developed chloracne. Thereafter
    the autoclaves were used for 2 years without incident but in 1958 a
    mechanic who conducted repair work on an autoclave subsequently
    developed chloracne (Hofmann, 1957; Goldmann, 1972). In 1968 and 1969
    the building containing the autoclaves was dismantled. Goldmann (1972,
    1973) conducted a study of the 42 workers exposed in this accident. In

    21 cases, the chloracne was preceded by a non-specific dermatitis and
    in two cases very persistent chronic conjunctivitis and blepharitis
    were observed; 14 cases also showed involvement of other organs. In
    four instances the liver was affected, and microscopic examination of
    the liver again showed a very characteristic grey pigment that did not
    stain positive for iron. A transient involvement of the myocardium was
    also noted. In five instances the upper respiratory tract was involved
    with tracheitis and bronchitis. There was one instance of haemorrhagic
    pleuritis and one instance of afebrile gingivitis and stomatitis. In
    a number of cases a high susceptibility to infection was noted,
    sometimes accompanied by a decrease in gamma-globulin. One worker died
    of pancreatitis, in seven cases the central nervous system was
    affected, three instances of toxic polyneuritis were recorded, and in
    two instances hearing, sense of smell, and taste were impaired. The
    child of one of these workers also developed chloracne, and in most of
    the workers active chloracne persisted for many years - in one
    instance for 18 years. Follow-up studies are described in section 9.4.

         Of particular interest is a study by Risse-Sundermann (1959).
    According to oral reports by the treating physician, all 24 members of
    a team working in a trichlorophenol operation became ill after the
    production process was switched to the pressurized phenol process in
    the spring and summer of 1954. Slightly different acneiform skin
    conditions appeared as symptoms of the toxic exposure. In addition,
    the patients suffered from dizziness, nausea, vomiting, lacrimation,
    burning of the eyes, difficulty in hearing, gastrointestinal spasms,
    intolerance to fatty foods, diarrhoea, jaundice, hepatitis (which was
    fatal in one case), and paresthesias and hyperesthesias, as well as
    extreme irritability. One patient became psychotic and committed
    suicide. In addition, some of the patients complained of impotence.

         Ten workers at this chemical factory were followed for five years
    by Risse-Sundermann (1959). In addition to the signs and symptoms
    mentioned above, she noticed swollen lymph glands and a considerable
    decrease in body weight. The patients underwent neurological
    examination, with no objective signs being observable. Of particular
    interest in this well documented study is the fact that in three
    patients the general symptoms (e.g., tiredness, depression, lack of
    appetite, stomach pains, sexual dysfunction) preceded that of the skin
    manifestations .

         Bauer et al. (1961) reported a study of workers affected by three
    different outbreaks of chloracne. In this study more than 100 workers
    were examined. Of these, 31 Hamburg workers had been exposed 5 years
    ealier. Nine were examined in detail and their symptoms tabulated.
    Initially, there was dermatitis and irritation of the face, sometimes
    accompanied by conjunctivitis, and followed by the gradual development
    of chloracne and patchy pigmentation of the skin. In some cases
    irritations of the mucous membranes of the face and upper respiratory
    tract, together with a persistent blepharoconjunctivitis, were also
    noted. In the follow-up study, a number of cases of liver injury were

    observed and, at liver biopsy, a typical grey pigment was observed in
    liver sections, which did not stain positive for iron. Viral hepatitis
    was suspected. In a few cases, chronic bronchitis and occasional
    myocardial damage were also observed. In all cases, fatigue was the
    main complaint and muscle weakness and muscle pain were described by
    the workers, particularly in the proximal muscles of the lower
    extremities. All nine also reported decreased libido. In a few
    instances, paresthesia and hyperesthesia or pronounced sensory
    neuropathy were observed, and minor circumscribed pareses were found.
    A psychovegetative syndrome occurred in most of the workers. Other
    signs recorded were: inability to concentrate, memory deficits, sleep
    disturbances, particularly increased somnolence, decreased drive, and
    alcohol intolerance. Psychological tests also showed abnormalities.

         Following the malfunction of a reaction vessel in northern Italy,
    in which 2,4,5-trichlorophenol was produced, the temperature in the
    vessel increased rapidly and an intense black vapour filled the
    work-room covering everything with a black deposit. Five workers
    engaged in clean-up operations developed chloracne (Hofmann &
    Meneghini, 1962). None of those involved in the clean-up exhibited any
    involvement of general systemic toxicity (even after 16 months) that
    could be related to exposure to the tar and soot. However, one
    15-year-old worker developed folliculitis and superficial nodular
    elements on the face a few days after initial exposure. A slow but
    progressive generalization of the dermatosis developed on the trunk,
    scalp, and lower extremities. An examination several months later
    revealed no damage to the renal or liver parenchyma. However, this
    worker was found to still suffer from outbursts of chloracne in 1980
    (Holmstedt, 1980).

         Duverne et al. (1964) reported a case that occurred at a plant at
    Lyon, France, where products that used 2,4,5-tri-chlorophenol as a
    starting material were manufactured. This worker developed chloracne
    as well as serofibrinous pleuritis.

         Ten workers also developed chloracne at a plant near Grenoble,
    France, which produced 2,4,5-trichlorophenol that served as the
    starting material for phenoxy pesticides and germicides for cosmetics.
    These workers showed symptoms of systemic poisoning similar to those
    reported by Goldmann (1972), and hepatic insufficiency with lipaemia
    and elevated serum cholesterol levels (Dugois & Colomb, 1956). Another
    accident resulting in TCDD exposure of workers occurred in the same
    factory in 1966 (Dugois et al., 1967).

         An exothermic reaction resulted in an explosion at a plant in
    Chesterfield, England, in 1968. The company made 2,4,5-trichlorophenol
    from tetrachlorobenzene and the explosion occurred during the process
    involving ethylene glycol and caustic soda under atmospheric pressure
    (Milnes, 1971). In this incident, 79 workers developed chloracne but
    there was no evidence of systemic illness (May, 1973). In 1971, 3
    years after the explosion, two workers who had not been involved in

    the explosion or its aftermath were employed as pipe-fitters at a new
    installation, away from the site of the explosion, to refit one of the
    cleaned tanks. They both developed severe chloracne, and the son of
    one of these workers and the wife of the other also developed this
    condition (Jensen & Walker, 1972). May (1973) cited two incidents
    involving explosions in a similar process. In the first incident,
    fatal injuries were recorded; in the second incident, all 50 exposed
    persons fell ill after 10 days and had liver injury.

         In the USA, an outbreak of chloracne occurred among workers
    manufacturing 2,4-dichlorophenoxyacetic acid and
    2,4,5-trichlorophenoxyacetic acid (Bleiberg et al., 1964); 29 workers
    developed chloracne and 11 of these had elevated urinary uroporphyrins
    and exhibited varying degrees of acquired porphyria cutanea tarda. At
    least one of these workers had abnormal liver-function tests and
    microscopic examination of a liver biopsy specimen showed parenchymal
    cell regeneration and haemofuscin pigment. Many of the workers with
    chloracne showed hyperpigmentation of the skin. A second study of the
    workers at this plant was conducted in 1969 by Poland et al. (1971).
    A total of 73 male employees were examined, and moderate to severe
    chloracne was found in 13 workers (18%), mild chloracne in 35 (48%),
    hyperpigmentation in 30, and uroporphyrinuria in 1. No definite
    systemic illness could be documented in these workers. Of those
    studied, 33 had been employed at the plant for 0-4 years, 10 for 4-8
    years, and 30 for more than 9 years. The mean duration of employment
    was 8.3  7.6 years (mean  1 SD). The trichlorophenol manufactured in
    this plant contained 10-25 mg TCDD/kg. Twenty-six of the workers seen
    by Bleiberg et al. (1964) were also seen in a follow-up study. Six
    months prior to the second survey (Poland et al., 1971), the
    manufacturing process was altered so that the 2,4,5-T produced
    contained less than 1 mg TCDD/kg.

         In 1964, in the USSR, many workers developed chloracne while
    engaged in producing 2,4,5-T. Production was then discontinued
    (Telegina & Bikbulatova, 1970).

         On 10 July 1976, an explosion occurred at the ICMESA plant at
    Meda, near Seveso, Italy, when 12 workers were present. All 176
    workers of the plant were examined 3 or 4 weeks after the accident.
    Chloracne was suspected in 1 of them; the others showed minor symptoms
    that could not be correlated with exposure. Alkaline phosphatase and
    delta-glutamyltransferase seemed slightly increased in 32 and 37
    cases, respectively, while five workers showed a reduction in their
    delta-aminolevulinic acid dehydratase blood levels, and three showed
    moderately increased urinary gamma-aminolevulinic acid (Zedda et al.,
    1976). Similar findings were reported by Fara et al. (1976) and
    Reggiani (1978). The follow-up of the general population is described
    in section 8.2.

    8.2  General Population Studies

    8.2.1  Missouri, USA

         Environmental exposures have occurred in a small area of
    Missouri, USA (Carter et al., 1975; Kimbrough et al., 1977; Kimbrough,
    1984). In the summer of 1971 many birds, rodents, cats, dogs, insects,
    and horses died after exposure in a horse arena in eastern Missouri.
    The incident followed the spraying of "waste oil" on the horse arena
    for dust control. Within 3 weeks of the spraying of this arena, two
    other arenas were sprayed. In all, 57 adult horses died, 26 abortions
    occurred among the horses at the most heavily exposed farm and many
    foals died soon after birth. At the time, the nature of the chemical
    that had caused the problem was unknown. The arenas were excavated and
    the contaminated dirt dumped at other sites. After many fruitless
    attempts to identify the cause of this outbreak, it was discovered in
    1973-1974 that the original soil from one of the arenas contained
    5600-6500 mg trichlorophenol/kg, 31.8-33.0 mg TCDD/kg, and 1350-1590
    mg polychlorinated biphenyls/kg. Because of this finding, the episode
    was reinvestigated. It was found that the salvage oil company that
    sprayed the three arenas routinely collected discarded motor oil and
    lubricants from over 2000 service stations in eastern Missouri and
    southwestern Illinois. It also collected, from various sources, a
    limited amount of used organic solvents such as transformer oils and
    other compounds. A company in southwestern Missouri was finally
    identified as a source of TCDD. This company had manufactured
    trichlorophenol as an intermediate for hexachlorophene. The production
    of 2,4,5-tri-chlorophenol had generated a distillate residue which was
    emptied once a week into a residue storage tank. Initially this
    chemical waste was collected and incinerated but, in 1971 when the
    trichlorophenol producer experienced a financial crisis, he arranged
    for the chemical wastes to be disposed of by a chemical supplier. The
    chemical supplier subcontracted the chemical waste disposal to the
    salvage oil dealer. The salvage oil dealer added the toxic chemical
    waste to his salvage oil storage tank, having collected a total of 18
    000 gallons. This material, mixed with salvage oil and other
    chemicals, was sprayed on the riding arenas and some of it was taken
    to re-refining companies. Soil samples from arenas where contaminated
    dirt had been dumped in 1974 contained trichlorophenol levels that
    ranged from 1.5-32.6 mg/kg, TCDD levels that ranged from 0.22-0.85
    mg/kg, and polychlorinated biphenyl levels that ranged from 10-25
    mg/kg.

         A 6-year-old girl, who had used one of the arenas for
    sandbox-like play in 1971, developed epistaxis, headache, diarrhoea
    and lethargy, haemorrhagic cystitis, and signs of pyelonephritis. She
    had an uneventful recovery. Three other females exposed to the same
    arena had recurring headaches, skin lesions, and polyarthralgias. Two

    3-year-old boys in another arena developed chloracne on the exposed
    skin surfaces which lasted for more than a year. Evaluation of the
    three female patients 5.4 years after exposure to TCDD-containing oil
    showed them to be in good health (Beale et al., 1977).

         A comprehensive medical examination of 154 residents exposed to
    TCDD, and 155 unexposed residents in similar type housing in eastern
    Missouri, revealed no consistent differences between the two groups.
    The examination included a medical history, physical examination,
    serum and urinary chemistries, and immunological and neurological
    tests. The findings may suggest that long-term TCDD exposure is
    associated with depressed cell-mediated immunity (decreased
    delayed-type hypersensitivity skin reactions to standard antigens)
    (Hoffman et al., 1986; Stehr et al., 1986). Urinary concentrations of
    glucaric acid were not significantly different between persons
    identified as being at high or low risk (Steinberg et al., 1985).

    8.2.2  Seveso, Italy

         The scientific follow-up on the population of Seveso, N. Italy,
    which had been accidentally exposed to TCDD in 1976 (see sections
    4.1.1 and 8.1), was guided by an international steering group headed
    by Professor M.A. Klingberg. The group completed its work in February
    1984 and concluded that "it is obvious that no clear-cut adverse
    health effects attributable to TCDD, besides chloracne, have been
    observed" (Regione Lombardia, 1984). A total of 193 people had
    displayed symptoms of chloracne, but at the beginning of 1984 only 20
    presented active symptoms. After 15-20 days exposure to TCDD soil
    levels of 270-1200 g/m2, there was a marked incidence of chloracne.
    No disturbance of biochemical functions were seen when the exposure
    had been limited to soil with TCDD levels at or below 30-70 g/m2.
    Later evaluations failed to confirm earlier findings of a decrease in
    motor nerve conduction velocity in some individuals. A significant
    increase in urinary glucaric acid levels, indicating increased
    microsomal enzyme activity, was found 3 years after exposure in 67
    exposed children, as compared to 86 non-exposed children (Ideo et al.,
    1982, 1985). The steering group found the data difficult to evaluate
    as analytical and individual biological variabilities were not
    explained (Regione Lombardia, 1984).

         Studies performed on the rate of spontaneous abortions and birth
    defects in the Seveso area do not allow any conclusions to be drawn
    (Tognoni & Bonaccorsi, 1982). The hypothesis that low exposure might
    cause pre-pregnancy or pregnancy effects that adversely affect the
    outcome was tested using several exposure models. The only finding was
    a slightly higher rate of haemangioma among newborns in the exposed
    group. However, this showed up only with one of the exposure models.
    It was considered doubtful that this was due to TCDD exposure (Regione
    Lombardia, 1984).

         Lymphocytes from inhabitants of Seveso were examined for
    chromosomal aberrations by Regianni (1980a,b) and Mottura et al.
    (1981). In 17 TCDD-exposed individuals examined within two weeks of
    the accident, no increase in chromosomal aberrations was observed
    (Regianni, 1980). In the abstract by Mottura et al. (1981),
    chromosomal aberration analysis was performed on subjects distributed
    into three classes: acute exposure, chronic exposure, and a control
    group of non-exposed subjects. No significant difference in the
    frequency of chromosomal aberrations in the three exposure categories
    was reported. Data on number of subjects, chromosomal aberrations, and
    exposure level and time were not given.

         Tenchini et al. (1983) published a comparative cytogenetic study
    on induced abortions from women exposed to TCDD after the Seveso
    accident, and in non-exposed subjects. Chromosome analysis was
    performed on maternal peripheral blood, placental and umbilical cord
    tissues, and fetal tissues. No significant differences were found in
    the level of chromosomal aberrations in the blood of placenta and
    umbilical cord from TCDD-exposed and non-exposed women. The exception
    was fetal samples from non-exposed women, in which a significant
    increase in chromosomal aberrations was obtained, possibly an artefact
    due to experimental techniques. The effect of TCDD on fetal
    chromosomes is therefore still unclear.

         Several epidemiological follow-up studies are continuing in and
    around Seveso.

    8.2.3  Viet Nam

         From 1960 to 1969 a mixture of 2,4-dichlorophenoxyacetic acid and
    2,4,5-trichlorophenoxyacetic acid (Agent Orange), which was
    contaminated with TCDD (concentrations ranging from 0.5 to 47 mg/kg)
    (Kearny et al., 1972), was sprayed over areas of Viet Nam as a
    defoliant. The spraying from 1960 to 1965 was minimal; in 1966 it
    covered slightly more than 800 000 acres, in 1967 almost 1.7 million
    acres, in 1968 over 1.3 million acres, and in 1969 1.2 million acres.
    Studies have been carried out since the early 1970s to ascertain
    whether the exposure of the general population in Viet Nam to this
    herbicide could have resulted in an increased incidence of birth
    defects. However, the results of such investigations have not been
    published in readily available peer-reviewed journals, making it
    difficult to assess the scientific significance of the findings. Such
    studies have been reviewed by Westing (1984) and Constable & Hatch
    (1985). Those studies reviewed indicate a range of effects including
    spontaneous abortions, infertility, and birth defects. However, there
    are marked deficiencies in experimental design in most, if not all,
    studies, including potential bias in the selection of populations,
    poor record-keeping of populations and biological effects, such as
    congenital malformations, and a lack of control over possible

    confounding factors. These deficiencies make it difficult, if not
    impossible, to use this body of data in assessing the human health
    risks from exposure to phenoxyherbicides contaminated with TCDD and
    other PCDDs.

         Tung (1973) reported an increased incidence of liver tumours in
    Viet Nam. From 1955 to 1961 there were 159 cases of liver cancer out
    of a total of 5492 cancer cases, and from 1962 to 1968, 791 out of a
    total of 7911 cancer cases. Van (1984) continued Tung's investigation.
    Previous exposure to herbicides of 21 male cases of primary liver
    cancer and 42 controls was ascertained. Six of the 21 cases and three
    of the controls had lived or worked in areas sprayed with herbicides
    or had moved there shortly after spraying ceased. Residence time
    varied from 8 to 77 months. There is a lack of information on
    confounding factors and there was a chance for bias in these studies.
    In general, the possibility of exposure to multiple chemicals and the
    short latency period noted make the study by Van (1984) of little
    value for assessing risk (IARC, 1986).

         In 1979, the United States Air Force (USAF) initiated an
    epidemiological study into the possible health effects from chemical
    exposure of Air Force personnel who conducted aerial dissemination of
    herbicide in Viet Nam (Operation Ranch Hand) (Lathrop et al., 1984).
    The purpose of this investigation was to determine whether long-term
    health effects exist and can be attributed to occupational exposure to
    herbicides. This study used a matched cohort design in a
    non-concurrent prospective setting, incorporating mortality,
    morbidity, and follow-up studies. The report presented the results of
    health information on 2706 Ranch Handers and comparison individuals
    obtained by questionnaire and 2269 Ranch Handers and comparison
    individuals undergoing an extensive physical examination. It was
    concluded that there was insufficient evidence to support a cause and
    effect relationship between herbicide exposure and adverse health in
    the Ranch Hand group at this time. The study disclosed numerous
    medical findings, mostly of a minor or undetermined nature, that
    require detailed follow-up.

         In a study of 15 soldiers in Australia exposed to Agent Orange,
    no increases in structural chromosomal aberrations or sister chromatid
    exchanges were noticed, compared to a control group of 8 subjects
    (Mulcahy, 1980). In 1980 the Australian Commonwealth Institute of
    Health agreed to conduct a series of scientific investigations into
    the health of Viet Nam veterans and their families. After
    considerating the most appropriate study programme, it was decided in
    1981 to conduct, as part of that programme, a case-control study of
    congenital anomalies and Viet Nam service (Donovan et al., 1984). The
    report is largely negative, as is that of Erickson et al. (1984) which
    reported a similar study of American veterans.

    
    Table 64. Signs and symptoms reported in association with human exposure
    to TCDD or mixtures containing TCDD
                                                                                  
    A. Skin Manifestations                       C. Neurological Effects

    1. Chloracne                                 1. Sexual dysfunction
    2. Hyperkeratosis                            2. Headache
    3. Hyperpigmentation                         3. Neuropathy
    4. Hirsutism                                 4. Sight disturbance
    5. Elastosis                                 5. Loss of hearing, taste,
                                                    and smell


    B. Systemic Effects                          D. Psychiatric Effects

    1. Mild fibrosis of liver                    1. Sleep disturbance
    2. Raised transaminase values                2. Depression
       in blood                                  3. Loss of energy and drive
    3. Hypercholesterolaemia                     4. Uncharacteristic bouts of anger
    4. Hypertriglyceridaemia
    5. Loss of appetite and weight loss
    6. Digestive disorders (intolerance to 
       alcohol or fatty food, flatulence,
       nausea, vomiting, diarrhoea)
    7. Muscular aches and pains, joint
       pain, lower extremity weakness
    8. Swollen lymph glands
    9. Cardiovascular, urinary tract,
       respiratory, and pancreatic disorders


                                                                               
    
    8.3  Signs and Symptoms in Humans Associated With TCDD Exposure

         Many signs and symptoms have been reported in studies of human
    exposures to PCDDs, both occupationally and from the general
    environment. These have been compiled from the various studies and are
    shown in Table 64.

    8.3.1  Skin manifestations

         Chloracne is a sign of exposure to several chlorinated cyclic
    organic compounds, the most potent being TCDD. Chloracne thus may
    serve as a marker of such exposure. The most distinctive lesion in
    chloracne is the so-called cyst, a skin-coloured elevation that may
    measure from 1 mm to 1 cm in diameter, with a central opening that may
    be difficult to detect. Comedones that contain black or
    black-appearing material in their openings are also present. There may

    be a secondary inflammatory reaction, melanosis, and hyperkeratosis,
    and these skin changes may be preceded by a "cable rash" or "cable
    itch". These skin lesions resemble photosensitivity reactions and the
    bearers may suffer severe pruritus. Microscopic examination of the
    skin lesions shows marked dilatation of the hair follicles which are
    filled with keratinous material, the sebaceous glands may be partly or
    completely atrophied and, occasionally, hyperplasia of these glands
    has also been reported. Hyperkeratosis and acanthosis of the
    surrounding epidermis usually accompany these lesions. Atrophy of the
    epithelium and thinning of the epithelial walls surrounding these
    keratinous cysts are observed at a later stage of the disease. If the
    follicular cysts rupture, foreign body granulomata may also be
    observed. Healing of these skin lesions usually results in deeply
    pitted scars. The distribution of chloracne is predominantly facial,
    affecting in particular the malar areas, the jaws, and the regions
    behind the ears. At times it may involve the ear canal and, with
    increasing severity, also the rest of the face and neck. In more
    extensive cases, the outer upper arms, neck, back, abdomen, outer
    thighs, and genitalia may also be involved (Crow, 1970).

         While the absence of chloracne does not absolutely rule out
    exposure to TCDD, it usually indicates that there has been no exposure
    to a toxic dose of the substance. "Toxic" is used here to indicate
    both systemic and local effects. Where there has been exposure to TCDD
    and chloracne has resulted, it is the only known clinical sign that
    persists for a long period of time, even for the remainder of the
    exposed person's life time. In a large group of people exposed to
    mixtures containing TCDD, the absence of chloracne usually indicates
    that exposure to a toxic dose was unlikely and also makes it unlikely
    that severe, persistent systemic disorders will result. 

         Hyperkeratosis is a fairly common phenomenon whereas
    hyperpigmentation and hirsutism are rare. It should be noted that
    hyperkeratosis is prominent in the exposed Seveso children who have no
    affected sebaceous glands. These glands develop only at puberty.

         Elastosis of the skin has been noted as a long-term effect of
    TCDD exposure.

    8.3.2  Systemic effects

         Liver effects following exposure to PCDDs have been diagnosed
    even by histological examination, and account for temporarily raised
    transaminases in blood, hypercholesteraemia, and
    hypertriglyceridaemia. Bauer et al. (1961) and Risse-Sundermann (1959)
    do not however exclude viral hepatitis as a cause of such findings in
    their patients exposed to TCDD. Loss of appetite, weight loss, and
    digestive disorders are common complaints from humans exposed to
    either TCDD itself, or to technical mixtures containing TCDD. Muscular
    aches and pain and weakness in extremities have been reported,
    particularly after exposure to technical mixtures containing TCDD.

    Swollen lymph nodes have also been reported, both after exposure to
    "pure" TCDD and to mixtures. The cardiovascular, urinary tract,
    respiratory, and pancreatic disorders reported are of doubtful
    significance with regard to a causal relationship to TCDD exposure.

         Porphyria cutanea tarda has been reported in two cases of
    occupational exposure where chlorinated organic compounds were
    manufactured in addition to trichlorophenol. These were the incidents
    at the factory of Diamond Alkali, Newark, New Jersey, USA, in 1956
    (Poland et al., 1971) and at Spolana, Czechoslovakia, between 1964 and
    1969 (Pazerova-Vejlupkova et al., 1981). The porphyria cutanea tarda
    observed in these cases was very unlikely to have been induced by
    exposure to TCDD but rather by exposure to other chlorinated organic
    compounds manufactured in these plants (Jones & Chelsky, 1986).

    8.3.3  Neurological effects

         Sexual dysfunction (lack of libido and impotence) has been
    reported after acute exposure to both "pure" TCDD and technical
    mixtures (Schulz, 1968). The frequency of its occurrence may have been
    underestimated to date. Headache is a frequent symptom after exposures
    to technical mixtures containing TCDD.

         Sensory neuropathy has been noted in many instances. Usually
    workers in the initial stages of exposure will complain of pains in
    their joints after they have very acute severe chloracne; however,
    there are usually no abnormal physical findings in the joints, but the
    complaints may continue. In early studies of workers affected by TCDD,
    no attempts were made to objectively measure the effects on the
    sensory nervous system. Tests have now been developed that evaluate
    sensory nerves and that can be used in future field studies. The nerve
    conduction tests, which primarily have been used so far, are actually
    not very useful to measure neuropathy. Differences in nerve conduction
    were shown among residents from Seveso, Italy, who had chloracne and
    those who did not (Richert, von, 1962; Fillipini et al., 1981).

         Sight disturbance may be related to alkaline exposure or to
    conjunctivitis related to effects on the glands of Meibom. Loss of
    hearing, taste, and smell have been reported in a few cases, but a
    causal relationship to TCDD exposure is doubtful.

    8.3.4  Psychiatric effects

         The symptoms have been listed in Table 64 in what is believed to
    be their order of frequency and degree of severity.

    8.4  Epidemiological Studies

         Signs and symptoms related to accidental exposure to TCDD are
    given in Table 64. However, it should be observed that all the
    accidents and occupational contamination concern exposure to a mixture
    of compounds where TCDD was only one component. In all cases, its
    concentration in the mixtures was unknown. Only two cases of
    intoxication with "pure" TCDD have been reported.

         The story of the discovery of TCDD is by now well documented
    (Holmstedt, 1980; Sandermann, 1984a,b). TCDD was synthetized in 1955.
    Four people were intoxicated, one co-worker severely so while drying
    crystals. In all cases, decreased libido was the first symptom,
    followed by other symptoms such as moderate to severe chloracne,
    sleeping difficulties, inability to concentrate, depression, and, in
    at least one case, swelling of the lymph nodes. In all cases, the
    signs and symptoms disappeared within a couple of years, with the
    exception of the chloracne in the most heavily exposed man.

         The second occasion of exposure to what one must assume to be
    pure TCDD is the one reported by Oliver (1975). The toxic effects on
    three young scientists who suffered "transient minimal exposure to
    TCDD" were described. Two of them suffered from typical chloracne.
    Delayed symptoms about two years after initial exposure occurred in
    two of the scientists. These symptoms were said to have included
    personality changes, other neurological disturbances, and hirsutism.
    All three scientists were found to have raised serum cholesterol
    levels, but no other biochemical disturbances and no porphyrinuria or
    liver damage were demonstrated. Whether the unusually delayed
    physiological effects were in fact due to the initial dioxin exposure
    is a question that was discussed by the author. Although conclusive
    evidence is lacking, it seems likely that these delayed effects were
    in fact due to dioxin intoxication. The conditions of exposure remain
    unexplained.

         Of the many cases of exposure reported in Table 63, only two (at
    Monsanto in 1949 and at BASF in 1953) have been adequately followed up
    epidemiologically with matched control groups.

         The workers of Monsanto, USA, have been investigated several
    times between 1949 and 1984. Immediately after the accident, Ashe &
    Suskind (1949) hospitalized and studied four cases of severe poisoning
    among the workers. These four workers were diagnosed as having
    chloracne, but by the time of examination these men had recovered from
    earlier symptoms of peripheral neuropathy. In 1950, a further
    examination of these four workers and two additional men revealed
    continued irritability, nervousness, and insomnia (Ashe & Suskind,
    1950). A consistent loss of libido and some impotence was reported.
    Further clinical examination revealed hepatomegaly, altered
    prothrombin times, and disturbed lipid metabolism.

         A further study of 36 workers from this plant was undertaken in
    1953 (Suskind et al., 1953). It was noted that even those who
    developed, to a moderate or severe degree, the skin eruptions, pains
    in back, dyspnoea, fatigue, nervousness, and decreased libido
    generally improved. Even those suffering the most severe cutaneous
    eruptions initially had only a few or no lesions in 1953.

         More recent studies on these workers are those of Zack & Suskind
    (1980) and Zack & Gaffey (1983). Zack and Gaffey reported on a
    121-member study cohort, with a presumptive high-peak exposure to TCDD
    base on chloracne occurrence, which was followed for mortality until
    1978. The entire cohort was traced; there were 32 deaths; 89 people
    were still alive. There was no excess in total mortality or in deaths
    from malignant neoplasms. The proportional mortality analysis of
    decedents according to exposure by 2,4,5-trichlorophenoxyacetic acid
    (2,4,5-T) indicated no unusual patterns of mortality. The proportional
    mortality ratio (PMR) for malignant neoplasms was low (PMR = 82) in
    the exposed group. Lung cancer was the only site among the malignant
    neoplasms for which the value was somewhat higher in the exposed
    group.

         The Monsanto workers were again examined in 1984 (Suskind &
    Herzberg, 1984). A clinical epidemiological study was conducted to
    determine the long-term health effects of workplace exposures during
    the process of manufacturing the herbicide 2,4,5-T, including
    contaminants such as TCDD. The population consisted of two cohorts,
    204 clearly exposed and 163 not exposed (controls). Among the exposed
    workers, clinical evidence of chloracne persisted in 55.7%. None of
    the controls experienced chloracne development. An association was
    found between the persistence of chloracne and the presence and
    severity of actinic elastosis of the skin. There was an association
    between exposure and the history of ulcers of the gastrointestinal
    tract. Pulmonary function values among those who were exposed and who
    currently smoked were lower than those who were not exposed and who
    currently smoked. No disturbances of sexual functions were found at
    this time after age adjustment. The data assembled in the study
    indicated no evidence of increased risk for cardio-vascular disease,
    hepatic disease, renal damage, or central or peripheral nervous system
    problems.

         Another selection of the population from the same plant has been
    examined by another group of epidemiologists (Moses et al., 1984).
    Since the degree of exposure was unknown to these investigators and
    since chloracne is generally considered a quite reliable indicator of
    heavy dioxin exposure, it was decided to use chloracne as a
    "surrogate" for exposure and to classify the study population by its
    presence or absence. It was recognized that those without chloracne,
    but with appropriate work-exposure history, might also have had TCDD
    exposure and were not therefore used as "unexposed controls".

    Chloracne was found in 52% of 226 workers in a 1979 cross-sectional
    survey at the plant where 2,4,5-T had been manufactured from 1948 to
    1969. Mean duration of residual chloracne was 26 years, and in 29
    subjects it had been present for 30 years. A significantly increased
    prevalence of abnormal gamma-glutamyl transpeptidase (GGT) and higher
    mean GGT were found in those with chloracne compared to those without.
    Although mean triglyceride values were higher in those with chloracne,
    the difference was not statistically significant. Neurological
    examination showed a statistically significant higher prevalence of
    abnormal sensory findings in those with chloracne. Increased
    prevalence of angina and reported myocardial infarction in those with
    chloracne was not significant when age-adjusted. Increased prevalence
    of reported sexual dysfunction and decreased libido in those with
    chloracne, compared to those without, was statistically significant
    after age adjustment. No differences were found between those with and
    without chloracne in serum cholesterol, total urinary porphyrins, or
    in reproductive outcomes. Exposure to TCDD in 2,4,5-T production may
    thus result in apparently permanent changes in the skin. Sensory
    changes in peripheral nerves and possible changes in liver metabolism
    in those with current or past chloracne are also suggested by these
    data. Based on worker histories, even severe acute toxicological
    effects of TCDD are reversible, or improve markedly over time. While
    the cross-sectional nature of this study, the low participation rate,
    and the highly select nature of the population limit the conclusions
    that can be drawn, it is unlikely that permanent, severe, and
    debilitating toxicological sequelae are inevitable after exposure to
    TCDD sufficient to produce chloracne. It must be noted, however, that
    individual susceptibility may make certain workers with heavy
    exposures more vulnerable.

         The exposure of workers at BASF in 1953 has been the subject of
    several reviews, the latest one being that of Thiess et al. (1982).
    Twenty-seven years after the accident that occurred in the BASF
    Ludwigshafen plant, a mortality study was undertaken of people exposed
    in the uncontrolled reaction which occurred during the trichlorophenol
    process. The follow-up was 100% successful and involved 74 people.
    Overall mortality (21 deaths) did not differ in this group from the
    rate expected in three external reference populations, or from that
    observed in two internal comparison groups, where 18-20 deaths were
    observed. Of the 21 deceased people, 7 had had cancer, compared with
    4.1 expected. In addition, two other cases of cancer (one bronchial
    carcinoma and one carcinoma of the prostate) were still alive at the
    time of writing. Three deaths due to stomach cancer at ages 64, 66,
    and 69 years, were found, compared with 0.6 expected from regional
    mortality data. One stomach cancer occurred among 148 individuals in
    the two comparison cohorts. The incidence of cancer in these workers
    was considerably greater than expected and cannot be explained only as
    a chance event. Of 74 people, 66 had severe chloracne or severe
    dermatitis. There is a possibility that some members of the BASF
    cohort were exposed to other unknown occupational hazards before or
    after the accident. However, the use of two internal comparison groups

    composed of matched controls from the same factory was designed to
    control for, as far as possible, other occupational exposures that
    could be important etiological or confounding factors. Because of the
    small size of the cohort and the small absolute number of deaths from
    any particular cause, the results of this study do not permit any
    definite conclusions concerning the carcinogenic effect of exposure.

         In comparison with the above-mentioned, well conducted long-term
    epidemiological studies, a host of other follow-up studies have been
    published, none of which used adequate controls. They are, therefore,
    of less value but will be briefly summarized here.

         Jirasek et al. (1973, 1976) and Pazderova et al. (1974, 1980,
    1981) examined 55 of a total of 80 workers who suffered intoxication
    during the manufacture of sodium pentachlorophenate and the sodium
    salt and butyl ester of 2,4,5-T. One worker died from severe acute
    intoxication at an early stage (Jirasek et al., 1976), and 76 workers
    developed chloracne. The following additional symptoms were found:
    porphyria cutanea tarda, disorders of the metabolism of lipids,
    porphyrins, and carbohydrates, and alteration of plasma proteins.
    Hepatic lesions were also present. Neurological and electromyographic
    (EMG) examinations revealed peripheral nerve changes in 17 people,
    first detected in 8 people during the second year of the study. A
    neurasthenic syndrome was also observed. The patients with porphyria
    cutanea tarda showed hyperpigmentation, hypertrichosis, and bullosis
    actinica mechanica. Porphyrin excretion in urine ranged from 172 to
    2230 g/24 h. Polyneuropathies, confirmed by EMG examination, were
    noted, predominantly in the lower extremities. In this outbreak, the
    disease was progressive during the first 2 years; subsequently the
    dermatological symptoms as well as the porphyric disease and the
    neurological disorders improved. The impaired lipid metabolism
    improved only very slowly.

         In this plant, the toxic substances were led off through the
    breathing zone of the workers. The concentrations of the chlorinated
    hydrocarbons in the air were never measured. Due to insufficient data,
    the real hygienic conditions at the work place could not be accurately
    reconstructed. The manufacturing of 2,4,5-T was halted permanently in
    1968 so that it was impossible to obtain the necessary information in
    an adequate manner. From 1959 to 1964, according to information from
    the plant, only sodium pentachlorophenate was manufactured. Not until
    1965 was the manufacture of sodium 2,4,5-T commenced on a pilot scale,
    and later the butyl ester of 2,4,5-T was also manufactured. After each
    year of production, something was always changed or modified in the
    process and technology so that actually there was never full-scale
    production in the true sense of the word. Many of the herbicides
    manufactured could not be found from the documentation (Pazderova et
    al., 1974). The uncertain mixture of compounds involved in the Spolana
    episode makes interpretation of signs and symptoms almost impossible.
    In all likelihood the porphyria observed was due to the
    hexachlorobenzene stated to be produced at this factory.

         Signs of disturbance in the porphyrin metabolism in workers
    manufacturing 2,4,5-T was also described by Poland et al. (1971).
    Chloracne was not correlated significantly with job location within
    the plant, duration of employment, or coproporphyrin excretion.
    Although 11 subjects with uroporphyria and at least three with overt
    porphyria cutanea tarda had been found in a study of the same plant
    six years earlier (Bleiberg et al., 1964), no clinical porphyria could
    be documented at the time of the second investigation, and only one
    worker had persistent uroporphyrinuria. Evidence of toxicity in other
    organ systems was markedly less than that reported in previous studies
    and in most instances there was no difference from normal populations.
    In all likelihood the porphyria cutanea tarda in this case, as in the
    study from Czechoslovakia, was due to a compound other than TCDD. This
    is corroborated by a recent re-evaluation of the literature (Jones &
    Chelsky, 1986).

         A study from northern Germany was published by Bauer et al.
    (1961). It is not clear where the cases orginated and only nine
    patients were studied in depth. A summary of the findings from these
    patients, and another person suffering from chloracne after
    occupational exposure to trichlorphenol, was reported by Kleu & Gltz
    (1971). These patients were followed for 15 years after exposure. The
    severity and types of symptoms varied in a dose-related manner. Major
    complaints were decreased sexual activity, muscular weakness, easy
    fatigability, irritability, and loss of appetite and memory. The
    authors concluded that a permanent defect had occurred, the late form
    of which resembled a cerebral involutionary syndrome, combined with
    mental depression and neurasthenia.

         A follow-up study of 11 of the 24 employees at Boehringer
    exhibiting skin lesions in 1955 was published by von Krause & Brassow
    (1978). Many continued to suffer from their earlier complaints. In
    seven of the eleven, nausea and intolerance to heavy fatty food was
    still common, and six men complained of alcohol intolerance. Although
    conjunctivitis had disappeared, chloracne was still clearly visible in
    most of the 11 subjects. Neurological problems were still severe in
    six of the workers.

         Ten years after the incident at Coalite in 1968 when 79 workers
    developed chloracne due to exposure to a chlorophenol-TCDD mixture, a
    study was undertaken to establish the state of health of the affected
    employees (46) remaining in the company's employment (May, 1982).
    Forty-one of the 46 employees participated. The opportunity was used
    to examine effects on mortality, morbidity, carcinogenesis,
    reproduction, teratogenicity, fetotoxicity, biochemistry, immunology,
    and genetic changes. Concurrently, two control groups were
    established, one with no dioxin exposure and the other with possible
    dioxin exposure. These groups were selected from within the works and
    matched the study group with respect to sex and age, but it was not

    possible to match them for occupation and social status. Half the
    affected subjects still had minor chloracne. Other than this finding,
    the authors concluded that the subjects had not been had been
    adversely affected in any way.

         Data on the mortality of workers at the Dow Chemical company have
    been provided in two papers (Cook et al., 1980; Ott et al., 1980). The
    first of these studies describes the mortality of a cohort of 61 males
    involved in the preparation of trichlorophenol. Forty-nine of these
    workers developed chloracne, presumably as a result of skin absorption
    of the process contaminant TCDD. Within the limitations posed by
    cohort size and length of follow-up, the exposure to chlorophenol-TCDD
    mixtures did not appear to have adversely affected mortality
    experience. Overall, four deaths occurred and 7.8 were expected. Of
    these, one death was due to cardiovascular disease (3.8 expected) and
    three deaths were attributed to cancer (1.6 expected). None of the
    findings was statistically significant. The second paper examined the
    mortality experience of 204 people exposed to 2,4,5-T during its
    manufacture from 1950 to 1971. Length of employment within the 2,4,5-T
    process area ranged from less than one year to a maximum of
    approximately ten years. Efforts to minimize TCDD contamination of the
    product resulted in non-detectable concentrations (less than 1 mg/kg)
    near the end of this period. Within the scope of this mortality
    survey, no adverse effects were observed with respect to occupational
    exposure to 2,4,5-T or to its feedstock, 2,4,5-trichlorophenol.

         Hardell and his co-workers in Sweden have conducted a series of
    case-control studies and reported an increased risk of soft-tissue
    sarcomas in men who were exposed to phenoxy herbicides and/or
    chlorophenols (Hardell & Sandsstrom, 1979; Hardell, 1981; Hardell et
    al., 1981; Hardell & Ericksson, 1981). These authors also reported a
    case-control study that suggested that phenoxyacetic acids and
    chlorophenols may predispose to Hodgkin's lymphoma (Hardell et al.,
    1981). The relative risk was higher for a group exposed to phenoxy
    herbicides including 2,4,5-T and chlorophenols, i.e., pesticides that
    may be contaminated with PCDDs and PCDFs. However, an increased risk
    was still found in a group exposed mainly to phenoxy herbicides such
    as MCPA, 2,4-D, mecoprop and dichloroprop, i.e., pesticides with low
    or no contamination with PCDDs and PCDFs.

         Analysis of fat levels of PCDDs and PCDFs in patients with soft
    tissue sarcomas and in controls failed to reveal any differences
    between the two groups (Nygren et al., 1986) (section 4.4.4.1).

         A cohort study on Swedish farmers and gardeners has been carried
    out recently (Wiklund & Holm, 1986). Despite the greatly increased use
    of phenoxyacetic acid herbicides from 1947 to 1970, no time-related
    increase in the relative risk of soft-tissue sarcoma was found in the
    cohort or in any of the subcohorts. The same was found by Hoar et al.

    (1986) although the latter study points to an increase in non-Hodgkin
    lymphoma. It should be noted that in all these studies the majority of
    the herbicides used did not contain TCDD.

         In follow-up studies of workers exposed to 2,4,5-T and its
    precursor 2,4,5-trichlorophenol (and therefore, presumably, also to
    TCDD), no excessive deaths due to any cause were registered (Cook et
    al., 1980; Ott et al., 1980; Zack & Suskind, 1980; Zack & Gaffey,
    1983).

         Honchar & Halperin (1981) merged the above four cohorts and found
    that three (2.9%) of the total 105 deaths were reported to be from
    soft-tissue sarcoma. Based on national statistics only 0.07% was
    expected to be due to this cause. Fingerhut et al. (1984) reviewed the
    employment records, medical and pathological reports, tissue
    specimens, and death certificates for these three cases and four
    additional cases of deaths from soft-tissue sarcomas in these and
    related cohorts reported by Cook (1981), Moses & Selikoff (1981), and
    Johnson et al. (1981). Three out of the seven cases had a record of
    chloracne and one of dermatitis. After review of the tissue specimens,
    five of the seven cases were diagnosed as soft-tissue sarcoma. The
    remaining two (which were among the three cases in the merged cohort
    of Honchar & Halperin (1981)) were found to be carcinoma. For three of
    the cases with confirmed soft-tissue sarcoma the exposure was not well
    documented, although an undocumented contact with 2,4,5-T,
    2,4,5-trichlorophenol, or TCDD could not be excluded.

    8.5  Human Experimental Studies

         Poiger & Schlatter (1986) studied a human volunteer after
    ingestion of a single dose of 1.14 g 3H-TCDD/kg body weight. The
    absorption from the intestine was > 87% and adipose tissue levels
    were 3.09 ( 0.05) and 2.85 ( 0.28) ng/kg after 13 and 69 days,
    respectively. The estimated half-life of TCDD was 2120 days.

         Gorski et al. (1984) calculated the half-lives of
    1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD and octaCDF to be about
    3.5, 3.6 and 2 years, respectively. The estimation was based on the
    analysis of fat tissue biopsies collected with an interval of 28
    months from one 14-year-old girl who for a period of about 2-3 years
    had been exposed to technical pentachlorophenol. Analysis was
    performed by gas chromatography with electron capture detection, and
    the isomers were confirmed by the use of several different packed and
    capillary columns.

    9.  TOXICOKINETICS OF PCDFs

    9.1  Uptake, Distribution, and Excretion

         Toxicokinetic data for TCDF and other PCDFs arise from iv
    injections or gastrointestinal exposure. There are no studies on
    exposure via the respiratory tract or via dermal application.

    9.1.1  Studies with 2,3,7,8-tetrachlorodibenzofuran (2,3,7,8-TCDF)

         Table 65 summarizes the distribution of radioactivity to the
    major tissue depots at various time-points after an iv injection of
    14C-TCDF into rats and mice. Similar studies in guinea-pigs and
    monkeys are summarized in Table 66. Tables 67 and 68 give the tissue
    distribution of 14C-TCDF in more detail for rats and mice,
    respectively.

         TCDF has been used for kinetic studies in the rat (Birnbaum et
    al., 1980), mouse (Decad et al., 1981a), guinea-pig (Decad et al.,
    1981b), and monkey (Birnbaum et al., 1981). A single iv dose of 30.6
    g 14C-TCDF/kg body weight was given to rats, mice, and monkeys,
    while the guinea-pigs received an iv dose of 6 g/kg body weight. The
    distribution of the radio-label was followed in tissues and excreta
    for 3 weeks in rats and monkeys, for 10 days in mice, and for 9 days
    in guinea-pigs. The distribution of radioactivity in the main tissues
    and excreta of the different species at some of the intervals studied
    is presented in Tables 65 and 66 along with the respective half-lives
    and LD50 values for TCDF. Radioactivity recovered from the tissues
    represented the parent compound, while radioactivity in faeces and
    urine represented metabolites of TCDF. In the faeces of guinea-pigs
    only the parent substance was present. Analysis by thin-layer
    chromatography revealed Rf values of 0.5 and 0.1 for metabolites of
    TCDF in faeces and urine as compared to an Rf of 0.8 for the parent
    compound.

         TCDF has a short half-life (2-4 days) and is quickly eliminated
    from the liver, both in the rat and the mouse (Birnbaum et al., 1980;
    Decad et al., 1981a). Elimination occurs rapidly also from the skin
    and muscle, whereas retention is longer in adipose tissues. The
    difference in the retention of TCDF in adipose tissues between C57Bl/6
    and DBA/2 mice may be explained by the fact that DBA/2 mice have
    substantially more adipose tissues than C57Bl/6 mice.

         The distribution of TCDF in the guinea-pig was different from
    that in the rat or mouse (Decad et al., 1981b). The maximum uptake in
    the liver occurred within one hour after dosing; thereafter the
    radioactivity was distributed in the fat and skin during the
    succeeding hours. After one day, as a result of loss of body fat, the
    radioactivity in adipose tissues was redistributed to the liver.
    Within 3 days after dosing there was no elimination of radioactivity
    from the liver and adipose tissues, whereas in the skin radioactivity
    decreased only slightly. The estimated half-life for TCDF in  the
    guinea-pig was more than 20 days.



    
    Table 65. LD50, whole-body half-life, and distribution of radioactivity (percentage of administered dose) at
    various intervals after an iv dose of 30.6 g 14C-TCDF/kg to ratsa and miceb
                                                                                                                             
                      Fisher 344 Rats                      C57BL/6 Mice                             DBA/2 Mice      

                3 hr        3 days     10 days     3 h          3 days       10 days      3 h          3 days        10 days
                                                                                                                            

    Liver       41.43.6     5.90.3      1.3      51.013.4    22.71.8      1.10.3     39.40.6     16.81.4       5.62.0
    Fat         10.01.0    11.12.3      1.8       6.01.6      2.92.1       ND          9.63.9     22.32.9       7.20.9
    Skin         6.60.3     1.20.3      0.5       3.60.7      3.01.1       ND          5.54.3      3.30.9         ND
    Muscle       5.90.4       0.3      < 0.3      7.52.8      1.50.9       ND         10.83.4      5.41.9        1.80.4
    Faeces                  63.10.6    > 85                      43.1       81.913.0                   27.7        55.84.8
    Urine                    2.00.4    < 6                        7.7       12.60.1                     9.2        19.94.6
                                                                                                                            

    Half-life               < 2b                                 2b                                     4b
    (days)
    LD50                  > 1000c                                 > 6000c,d
    (g/kg)
                                                                                                                            


    a Birnbaum et al. (1980).
    b Decad et al. (1981b).
    c Moore et al. (1976).
    d Moore et al. (1979).
    ND = not detectable.
    


        Table 66. LD50, whole-body half-life, and distribution of
    radioactivity (percentage of administered dose) at various intervals
    after an iv dose of 14C-TCDF in guinea-pigsa and monkeysb

                                                                               
                     Hartley guinea-pigs                  Rhesus monkeys
                3 h          3 days       9 days              21 days
                                                                               
    Liver       23.63.8     29.30.6     54.214.5           1.020.80
    Fat         31.40.7     56.97.6     21.811.6           3.662.83
    Skin        22.50.1     17.1+0.6    15.23.1            2.441.60
    Muscle                   15.64.5      8.83.0            1.550.14
    Faeces                    4.71.3       6.6                 42.9
    Urine                     2.30.4       6.6                  7.9
                                                                              

    Half-life                 > 20c                             8a
    (days)
    LD50                       > 5c < 10d                  1000d
    (g/kg)
                                                                              

    a  6 g/kg (Decad et al., 1981a).
    b  30.6 g/kg (Birnbaum et al., 1981).
    c  Moore et al. (1976).
    d  Moore et al. (1979).
    
        Table 67.   Tissue distribution of TCDF-derived radioactivity in
    Fisher 344 rats at 15 min, 3 h, and 24 h following a single iv dose
    of 30.6 g 14C-TCDF/kg b

                                                                               

    Tissue         Tissue content of 14C (% of dose/g tissue)a
                                                               
                   15 min         3 h            24 h
                                                                               
    Blood          0.120.04      0.040.01      0.030.01
    Liver          4.4 0.2       5.1 0.4       2.2 0.4
    Fat            0.200.03      0.440.07      0.640.11
    Muscle         0.250.01      0.060.00      0.300.02
    Skin           0.170.02      0.200.01      0.070.01
    Kidneys        0.670.04      0.170.03      0.080.01
    Adrenals       7.4 6.9       4.7 1.3       0.340.14
    Thymus         0.520.12      0.540.13      0.070.03
    Spleen         0.370.07      0.080.02      0.020.00
    Testes         0.090.01      0.090.01      0.060.02
    Brain          0.250.01      0.150.03      0.020.00
    Lungs          1.080.08      0.240.02      0.070.02
    Heart          0.660.03      0.110.00      0.020.02
                                                                               

    a  Mean  SD for three animals.
    b  From: Birnbaum et al. (1980).
    
         Based on data from three monkeys, the half-life for TCDF was
    calculated to be 8 days (Birnbaum et al., 1981). At the end of the
    study, more radioactivity remained in adipose tissues and skin than in
    the liver. The retention of TCDF in the liver of monkeys 21 days after
    dosing was comparable to that in the liver of the rat and C57Bl/6
    mouse 10 days after injection. Urinary elimination of radioactivity
    was a minor route when compared to faecal elimination both in the rat,
    mouse, and monkey, whereas in the guinea-pig these routes were of
    comparable importance (Birnbaum et al., 1980, 1981; Decad et al.,
    1981a,b). The cumulated excretion of radioactivity 3 days
    post-treatment amounted to approximately 64, 51, 11, and 7% in the
    rat, C57Bl/6-mouse, monkey, and guinea-pig, respectively.

         Against this background of data on tissue distributions,
    half-lives, and LD50 values of TCDF in the rat, guinea-pig, and
    monkey, Birnbaum et al. (1980, 1981) concluded that TCDF, measured as
    excreted radioactivity, is metabolized to less toxic compounds and
    that animal species with a high capacity to metabolize TCDF are more
    resistant to its acute toxicity. This conclusion was considered
    applicable also to the mouse (Decad et al., 1981a). Based on the same
    data King et al. (1983) produced a pharmacokinetic model for TCDF in
    rats, mice, and monkeys. However, there are objections to this
    comprehensive conclusion. First, the kinetic studies on guinea-pigs
    (Decad et al., 1981b) were (for analytical reasons) carried out with
    such a high dose of TCDF that all of  the animals showed marked signs
    of toxicity, even within 3 days. After 9 days all the animals were
    killed due to toxic  symptoms. It is not advisable to draw any
    conclusions on normal kinetic behaviour from data obtained on dying
    animals with their abnormal metabolism and physiology. As far as the 
    kinetic data from the monkey are concerned, the conclusions were based
    on a single time-point, and the number of animals in that study was
    also very limited (Birnbaum et al., 1981).



    
    Table 68. Tissue distributiona of TCDF-derived radioactivity in C57Bl/6 and DBA/2 mice at 15 min, 3 h, and 24 h after
    a single iv dose of 30.6 g 14C-TCDF/kg b
                                                                                                                             

                                                   Tissue content of 14C (% of dose/g tissue)a
                                   15 min                         3 h                           24 h           
           Tissue            C57Bl/6        DBA/2          C57Bl/6        DBA/2          C57Bl/6        DBA/2
                                                                                                                             
           Blood                1.1            ND           0.60.3        0.220.04        ND           0.20.0
           Liver             28.04.2       30.72.8       39.32.8       38.0 3.8      25.34.2       19.21.9
           Adipose            2.60.4        2.71.0        3.71.4        4.9 0.1       6.11.2        6.10.4
           Muscle             1.30.1        1.60.2        0.70.2        0.9 0.3       0.30.1        0.80.1
           Skin               2.20.1        2.30.7        1.40.2        2.7 0.6       1.30.7        2.70.6
           Kidneys            3.40.3        4.10.5        1.10.3        1.3 0.3       0.60.1        0.70.2
           Adrenals          18.84.9          ND           6.95.4          9.2            6.5            ND
           Thymus             3.92.4        3.01.9        2.00.7        4.7 3.2       0.30.2        2.50.6
           Spleen             1.40.1        1.70.3        0.80.1        0.5 0.1       0.20.1         0.27
           Testes             0.40.1        0.60.1        0.50.2        1.0 0.2       0.20.2        0.40.3
           Brain              1.70.5        2.40.3        0.80.1        1.3 0.3       0.20.3        2.72.7
           Lungs              6.70.5        8.20.7        2.40.9        3.1 1.5       0.40.2        0.80.4
           Heart              2.11.3        3.42.2        0.60.3        0.8 0.4       0.20.1        0.30.1
                                                                                                                             

    a  Mean  SD for three animals.
    b  From:  Decad et al. (1981b).
    ND = below limit for accurate detection.
    


         Ioannou et al. (1983) calculated a whole-body half-life of
    approximately 40 days for a non-toxic dose of TCDF in young  male
    Hartley guinea-pigs. Their estimation was based on the distribution of
    TCDF-derived radioactivity in liver, adipose tissue, skin, and muscle
    in three animals 36 days after a single oral dose of 4 mg TCDF/kg body
    weight and on certain approximations obtained from a previous study
    (Decad et al., 1981b). Failure to demonstrate a correlation between
    degree of bioaccumulation and lethality of TCDF in this study may be 
    due partly to the calculations of body burden based on the uncertain
    estimate of the 40 days half-life for TCDF, which may not be valid for
    both toxic and non-toxic doses of TCDF.

    9.1.2  Studies with other PCDFs

         Young male Wistar rats absorbed approximately 68% of a single
    oral dose of 1.0 mg 2,3,4,7,8-pentaCDF/kg body weight given in salad
    oil (Yoshimura et al., 1986). The daily faecal excretion was about
    0.1% of the administered dose/day, whereas no 2,3,4,7,8-pentaCDF was
    detected in urine. Four weeks after dosing the retention of
    2,3,4,7,8-pentaCDF in the liver was 48.8% of the dose. The addition of
    5% of activated charcoal beads to the diet, one week after dosing and
    throughout the study, increased the faecal elimination of
    2,3,4,7,8-penta CDF about 3-fold, but had no effect on urinary
    elimination. Both the liver and extrahepatic tissues, except the
    kidney, from rats on basal diet supplemented with activated charcoal
    beads had lower levels of 2,3,4,7,8-pentaCDF than rats on basal diet
    only.

         Yoshihara et al. (1981) administered single ip injections of 13
    individual PCDF congeners (at 1, 5, or 10 g/kg) to young male Wistar
    rats, and retention of the respective isomers in the liver was
    determined 5 days later. The great variation observed in the hepatic
    accumulation of the various isomers seemed to depend on the position
    as well as the number of chlorine atoms substituted. Isomers having
    vicinal hydrogens were accumulated to a lesser degree, although three
    of the six isomers having no vicinal hydrogens (1,3,6,8-tetraCDF,
    TCDF, and 1,2,4,6,8-pentaCDF) also showed low accumulation. The isomer
    most highly accumulated was 2,3,4,7,8-pentaCDF, more than 65% of the
    dose being retained, whereas only 3.8% of TCDF, which is equally
    potent biologically, was retained. These results would imply that
    there is no relationship between hepatic distribution of PCDFs and
    their potential for acute toxicity. All animals in this study showed
    toxic symptoms and liver microsomal AHH activity was strongly induced,
    except in the cases of the following isomers: 2,8-diCDF,
    1,2,7,8-tetraCDF, 1,3,6,7-tetraCDF, 1,3,6,8-tetraCDF,
    1,2,4,6,8-pentaCDF. It is important to take this into consideration
    when judging the kinetic data. A mixture of 14% 1,2,7,8-tetraCDF, 35%
    TCDF, 1% 1,2,4,7,8-pentaCDF, 49% 1,2,3,7,8-pentaCDF, 1%
    2,3,4,7,8-pentaCDF, and 1% hexaCDF was administered as a single ip

    dose of 10 mg PCDF/kg body weight to young male Wistar rats (Kuroki et
    al., 1980). The retention of the isomers in the liver, 5 days
    post-treatment, showed good agreement with the results of Yoshihara et
    al. (1981).

         Based on the purification of three isoenzymes of cytochrome P-450
    and the recovery of 14C-radioactivity from the hepatic microsomes of
    Wistar rats treated with 14C-2,3,4,7,8-pentaCDF (single ip dose of
    1 mg/kg body weight, 5 days previously), Kuroki et al. (1986)
    suggested that one of these isoenzymes, P-448 H, functions as the
    storage site of 2,3,4,7,8-pentaCDF in the rat liver.

         Fly ash and crude or purified toluene extracts of PCDD- and
    PCDF-containing fly ash from a municipal incinerator (Zaanstad, The
    Netherlands) were mixed with ordinary laboratory diet for rats (van
    den Berg et al., 1983). Small portions (2 g) of these diets were fed
    to male Wistar rats (300 g) every 24 h for 19 days, at which time the
    animals were sacrificed. The levels of tetra-. penta-, and
    hexa-chlorinated PCDDs and PCDFs in samples of liver and adipose
    tissue from these rats were determined. Rats fed the fly
    ash-containing diet stored PCDDs and PCDFs in their livers at
    concentrations which were at least 3 to 5 times lower than those of
    rats fed with comparable amounts of fly ash extracts. For the
    pentaCDD, hexaCDF, and hexaCDD isomers these concentrations were
    approximately 10-20 times lower. Generally PCDFs had a higher
    retention in the liver of rats than the corresponding PCDDs. In the
    adipose tissue of rats fed with fly ash extracts, retention was higher
    for penta- and hexaCDDs than for the corresponding PCDFs.

         In a later study, male Wistar rats (275 g) were fed a diet
    containing the same fly ash, pretreated with 2.5% HCl (van den Berg et
    al., 1986a). A control group received standard diet. All congeners
    retained in the livers of the rats had a 2,3,7,8-chlorine substitution
    pattern. With the exception of 2,3,4,7,8-pentaCDF and
    2,3,4,6,7,8-hexaCDF, the retention for each congener was below 10% of
    the dose. The retention percentages of the various congeners in the
    liver were almost equal at all time-points studied (34, 59, and 99
    days), thus indicating a long half-life of these congeners in the
    liver of the rat.

         A mixture of two tetraCDFs, four pentaCDFs, and four hexaCDFs,
    was given as a single ip injection of 500 g to male ICR mice (Morita
    & Oishi, 1977). The distribution patterns of the isomers in various
    tissues were followed for up to 8 weeks. Analyses were performed with
    GC (with electron capture detection) and isomers were identified by
    peak number only. PCDFs were mainly located in the liver, spleen, and
    fat tissues, but low to minimal amounts were found also in the kidney,
    testes, lungs, heart, and brain. The GC patterns of liver samples
    changed markedly with time, in contrast to those of the other tissues,
    including fat, where the GC patterns remained similar throughout the

    study. Most isomers with shorter retention times were readily absorbed
    and then rapidly disappeared from the liver. Isomers with longer
    retention times were slowly absorbed, and thus appeared later and
    persisted longer in the liver. If the mixture had been administered
    orally, those isomers with long retention times might have passed the
    gastrointestinal tract with very low absorption.

         The hepatic retention of PCDDs and PCDFs after dietary intake of
    the above-mentioned HCl-pretreated fly ash was studied in male Golden
    Syrian hamsters (van den Berg et al., 1986b). The livers were analyzed
    for tetra-, penta-, and hexaCDDs and -CDFs after feeding the diet,
    which contained 25% fly ash. No detectable hepatic retention was
    observed after 34 days. The highest retention after 95 days was 8.4%
    for 2,3,4,7,8-pentaCDF, but the retention was generally below 5% of
    the total dose. With the exception of 2,3,4,6,7- pentaCDF, only
    2,3,7,8-substituted PCDDs and PCDFs were retained. Constant relative
    concentrations were found for the 2,3,7,8-substituted PCDDs and PCDFs
    at the time-points studied.

         In studies by Firestone et al. (1979), three lactating Holstein
    cows received commercial grade pentachlorophenol orally by gelatine
    capsule at a dose rate of 10 mg/kg body weight twice daily for 10 days
    and once daily for the following 60 days. One cow served as a control
    and received gelatine capsules containing only ground corn. The
    pentachlorophenol composite used contained ten PCDD congeners (0.1-
    690 mg/kg) and eight PCDF congeners (0.9-130 mg/kg). Faeces collected
    on day 28 of the treatment period contained three hexaCDDs (0.05-0.63
    g/kg), two heptaCDDs (21.3-33.1 g/kg), and octaCDD (290-429 g/kg).
    Faeces also contained hexa-, hepta-, and octaCDF. Milk, body fat, and
    blood contained only three of the PCDD congeners present in the
    pentachlorophenol composite, namely 1,2,3,6,7,8-hexaCDD,
    1,2,3,4,6,7,8-heptaCDD, and octaCDD. Milk samples also contained
    hexa-, hepta-, and octaCDF. The average concentrations of
    1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, octaCDD, and octaCDF in
    the composite milk fat at the end of the treatment period were 20, 40,
    25, and 2 mg/kg, respectively. Similar concentrations were found in
    body (shoulder) fat at the end of the treatment period (13, 24, and 32
    mg/kg, respectively, for 1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD,
    and octaCDD). Levels of dioxins in the blood were about 1000 times
    below the values in milk or body fat. The average daily excretion of
    1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, and octaCDD in the milk
    during days 40 to 70 of treatment was about 20, 40, and 23 mg
    (corresponding, respectively, to 33, 3, and 0.6% of the daily intake
    of PCDDs). One hundred days after cessation of treatment the average
    levels of 1,2,3,6,7,8-hexaCDD, 1,2,3,4,6,7,8-heptaCDD, and octaCDD in
    shoulder fat and milk fat were 2.5, 6.6, 5.6 mg/kg and 4.3, 6.9, 3.0
    mg/kg, respectively.



    
    Table 69. Levels of PCDFs in the liver of dams and in fetuses and offspring after oral administration of PCDFs to
    mice for 18 days during pregnancyf
                                                                                                                             

    PCDF                         Total intake           % of PCDF intake in:a     Total intake     % of PCDF intake in:b
    congener                     of PCDFs by                                      of PCDFs by
                                 dams killed                                      dams killed
                                 on day 18 of           liver                     2 weeks after    liver       offspring
                                 pregnancy (g)d        of dam   fetus            delivery (g)    of dam      (week)
                                                                                                               1     2
                                                                                                                             

              tetraCDFc           1.40.06d              5.4     ND                1.60.10         ND         ND    ND
              tetraCDFc           8.10.32               ND      ND                9.00.57         ND         ND    ND
      2,3,7,8-tetraCDF           11.49.46               5.5     0.007            12.50.79         0.03       0.05  Te

              pentaCDFc           2.90.12               5.7     ND                3.10.20         4.0        0.10  0.27
              pentaCDFc          13.20.53               3.9     0.004            14.50.91         0.1        0.03  Te
    2,3,4,7,8-pentaCDF           4.90.20               14.5     ND                5.40.34        10.0        0.29  0.89

              hexaCDFc            1.40.06              10.7     ND                1.60.10         6.4        0.28  0.76

    Total                        43.21.73               5.2     0.003            47.73.0          1.6        0.07  0.14
                                                                                                                             

    a  Nine dams in group killed on day 18 of pregnancy.
    b  Ten dams in group killed on day 14 after delivery.
    c  Specific isomer not determined.
    d  Mean  SEM.
    e  T = 0.01-0.1 g/kg total congener.
    f  From: Nagayama et al. (1980).
    ND = not detected.
    


    9.2  Metabolic Transformation

         Thirteen chlorinated compounds were detected in bile collected
    for 48 h from female Sprague Dawley rats given a single oral dose of
    678 g TCDF (79.4% pure)/kg body weight (Poiger et al., 1984). The
    four major metabolites considered to originate from TCDF were
    trichloromethoxy-dibenzofuran, two trichlorodimethoxy-dibenzofurans,
    and tetrachloromethoxy-dibenzofuran. The remaining nine metabolites,
    detected in minute amounts, originated most likely from contaminating
    PCDFs (1% triCDF, 8.4% tetraCDFs, 11.2% pentaCDFs).

         Metabolites of 2-monoCDF, 2,8-diCDF, 2,3,8-triCDF, and octaCDF
    were determined in the urine, faeces, fat, and liver of male Wistar
    rats given single oral doses of 250 mg/kg body weight of the
    respective isomers (Veerkamp et al., 1981). Analyses were performed
    with GC-MS. No metabolites in any samples were found in rats given
    octaCDF. Monohydroxy and dihydroxy derivatives were obtained with all
    other isomers, whereas sulfur-containing metabolites were detected
    only with the monoCDF and diCDF. Metabolites from 2-monoCDF and 2,3,8-
    triCDF were found in urine and faeces only, but with 2,8-diCDF
    metabolites appeared also in the tissues.

    9.3  Transfer Via Placenta and/or Milk

         Nagayama et al. (1980) studied the transport of a mixture of
    PCDFs into the placenta and milk in the mouse (Table 69). A diet
    containing 0.6 mg PCDFs/kg (48% tetraCDFs, 49% pentaCDFs, and 3%
    hexaCDFs) was given for 18 days after mating. Nine dams were killed on
    day 18 of pregnancy and 10 on day 14 after delivery. After giving
    birth, the mothers were fed a diet free of PCDFs. The placental
    transport, calculated from the amount of PCDFs in the neonates, was
    about 0.003% of the administered dose. The isomers that remained in
    the tissues were TCDF and 2,3,4,7,8-pentaCDF. While the levels of
    PCDFs in the mothers dropped from 5.2% to 1.6% of the total intake,
    the whole-body levels in the sucklings increased from 0.003% at the
    time of birth to 0.07% after one week and to 0.14% of the total intake
    after 2 weeks. TCDF and 2,3,4,7,8- pentaCDF were the dominant species
    in both the mothers and the pups. To study the transport through milk
    only, the same PCDF-containing diet was given to pregnant rats for 14
    days from day 18 after mating, including the lactation period
    (Nagayama et al., 1980) (Table 70). After 14 days 5.1% of the total
    intake was found in the liver of the mother. The offspring contained
    0.3% and 1.2% of the dam's intake after 1 and 2 weeks, respectively.
    The dominant isomers recovered in the offspring were TCDF,
    2,3,4,7,8-pentaCDF, and one unidentified pentaCDF, i.e., the same
    isomers found in the largest amounts in the mother's liver. The data
    demonstrated that the amounts of PCDFs transferred through milk were
    much larger than the amounts transferred across the placenta.

    
    Table 70. Levels of PCDFs in the liver of mouse dams and in offspring after oral 
              administration of PCDFs to dams for 14 days following deliverye

                                                                               

    PCDF                     Total                  % of PCDF intake in:a
    congener                 intake
                             of PCDF        liver of dam   offspring (week)
                             (g)c                         1              2
                                                                               

              tetraCDFb       2.30.13         6.9         ND             ND
              tetraCDFb      12.90.71         ND          ND             ND
      2,3,7,8-tetraCDF       17.93.10         5.6         0.5            1.4

              pentaCDFb       4.40.24         6.7         Td             0.6
              pentaCDFb      21.01.14         3.7         0.4            1.6
    2,3,4,7,8-pentaCDF        7.70.42        13.8         0.4            2.2

              hexaCDFb        2.30.13         7.9         Td             1.5

    Total                    68.33.75         5.1         0.3            1.2
                                                                               

    a  Ten dams in the group.
    b  Specific isomer not known.
    c  Mean  SEM.
    d  T = 0.01-0.1 g/kg total congener.
    e  From: Nagayama et al. (1980).
    ND = not detected.
    
         Weber & Birnbaum (1985) studied the distribution and placental
    transfer of a single oral dose of 800 g 14C- TCDF (0.0485
    Ci/g)/kg body weight to pregnant C57Bl/6 mice on gestation day 11.
    Embryo mortality on gestation days 12 to 14 was in the range 7.9 to
    17.8%. No detectable radioactivity was found in the embryos whereas
    about 0.01% of the radioactive dose was contained in the placenta. The
    hepatic radioactivity in the dams decreased rapidly from 30.0% of the
    dose on gestation day 12 to 12.1% of the dose on gestation day 14. The
    cumulative urinary and faecal excretion from gestation day 12 to 14
    were 5.4 and 80.1% of the administered dose, respectively.

    10.  EFFECTS OF PCDFs ON ANIMALS

    10.1  Acute Toxicity

         Single oral LD50 values for TCDF in three species are listed in
    Table 71.

    10.1.1  Studies on rats

         No histological changes associated with TCDF toxicity could be
    observed in rats at oral doses up to 1000 g TCDF/kg body weight
    (Moore et al., 1976). These preliminary results, which also mentioned
    that only mild toxicological changes occurred in rats at 6000 g
    TCDF/kg body weight, have not been presented in a final report.

         Intravenous administration of 30.6 g TCDF/kg body weight to male
    Fisher rats has been shown to cause listlessness, excessive hair loss,
    and decreased weight gain 2 days post-treatment. These adverse effects
    were reversible and 3 weeks after dosing the animals appeared healthy
    with normal body weight. There were no signs of thymic or splenic
    atrophy or of liver hypertrophy (Birnbaum et al., 1980).

         Single ip injection of 1 or 10 mg/kg body weight of nine
    individual PCDF isomers, with at least three chlorines in the lateral
    positions, to male Wistar rats produced thymus atrophy and liver
    hypertrophy 5 days post-treatment. Five other congeners having no more
    than two chlorine atoms in the lateral positions did not cause any
    effects on the thymus or liver within the same dose range (Yoshihara
    et al., 1981).

         The ability of 15 individual PCDFs to affect body weight gain and
    thymic atrophy in immature male Wistar rats was investigated 14 days
    after a single ip injection (Ganon et al., 1985). The ED50 values
    for both effects were estimated for each congener (Table 61).

    10.1.2  Studies on mice

         Moore et al. (1976) failed to establish a lethal dose for TCDF in
    C57Bl/6 mice after giving single oral doses of up to 6000 g/kg body
    weight with an observation period of 30 days. However, there was a
    transient depression in body weight gain, thymic involution, and mild
    hepatotoxicity when 6000 g TCDF/kg was given subcutaneously. Poland
    & Glover (1980) found TCDF-induced thymus atrophy in C57Bl/6 mice 5
    days after a single ip dose of 3 x 10-7 mol/kg body weight.

         Single doses of 100 to 1000 g TCDF/kg body weight to pregnant
    C57Bl/6 mice on gestation days 10 to 13 produced no toxic effects on
    the dams within the time studied (Hassoun et al., 1984a; Weber et al.,
    1984).

         A mixture of two tetraCDFs, four pentaCDFs and four hexaCDFs
    given as a single ip dose of 500 mg to ICR mice produced no
    deathswithin 8 weeks (Morita & Oishi, 1977). CF-1 mice given a PCDF
    mixture containing 42% tetraCDFs, 54% pentaCDFs, and 4% hexaCDFs as a
    single oral (10 to 1000 mg/kg body weight), sc (10 to 200 mg/kg), or
    ip (10 to 100 mg/kg) dose developed no toxic signs during the first
    week, although a modest weight loss was noted (Nishizumi, 1978). The
    first deaths occurred 8 days after an oral dose of 1000 mg/kg, 5 weeks
    after a sc dose of 200 mg/kg, and 11 days after an ip dose of 100
    mg/kg. The oral LD50 (30 day) was 184 mg/kg for males and 414 mg/kg
    for females. Hepatomegaly and thymus atrophy were consistent findings
    in mice that died. In surviving mice on high dosages, the liver
    exhibited small necrotic foci accompanied by cellular infiltrates. The
    hepatic lesions occurred in the centrilobular area, and enlarged
    hepatocytes containing foamy cytoplasm, increased numbers of lipid
    droplets, and proliferation of smooth endoplasmatic reticulum were
    also seen.

    10.1.3  Studies on guinea-pigs

         In studies by Moore et al. (1979), the patterns of toxicity were
    similar for TCDF, 2,3,4,7,8-pentaCDF, and 2,3,7,8-
    tetrabromodibenzofuran when given to young Hartley guinea-pigs. The
    single oral LD50 was 5-10 g/kg body weight for all three isomers,
    and the time to death ranged from 8 to 26 days. Overt signs at lethal
    doses were immediate and progressive weight loss, rough soiled hair
    coat, listlessness, and dehydration. Similar symptoms appeared 3 days
    after an iv injection of 6 g TCDF/kg body weight (Decad et al.,
    1981b). At necropsy lack of body fat and reduced body mass and thymus
    weight were found. Histological findings were primarily associated
    with the depletion of lymphoid cells in the thymic cortex, but
    hypocellularity of bone marrow and hyperplasia in epithelial cells of
    the renal pelvis, ureter, and urinary bladder were also observed.
    Liver lesions were not observed. Surviving animals showed mild thymic
    lymphoid hypoplasia only. Sublethal doses resulted in decreased body
    weight gain.

         In a study by Ioannou et al. (1983), all three adult male Hartley
    guinea-pigs survived a single oral dose of 6 g TCDF/kg body weight
    for at least 17 days. At 10 and 15 g TCDF/kg body weight, deaths
    occurred on days 15-39 (two animals were sacrificed on day 17) and
    13-20, respectively. The acute oral toxicity of soot (and of benzene
    extracts of the soot) containing PCDDs and PCDFs from a PCB-containing
    transformer fire (Binghamton, New York, USA) were studied in female
    Hartley guinea-pigs (Silkworth et al., 1982). As discussed in section
    8.1.1.1, toxicities were noted at 100 and 500 mg/kg body weight of
    soot, but not when 1 and 10 mg/kg were administered.



    
    Table 71.  Single lethal dose values for TCDFa
                                                                                                                             
         Species/strain      Sex/No    Age or    Dose                Duration       LD50           Time to
                                       weight    tested              of study       (g/kg         death
                                                 (g/kgc             (days)         body weight)   (days)
                                                 body weight)
                                                                                                                             
         Mice
         (C57Bl/6)           M/8       6 weeks       0                 30           > 6000         not
                                                   400                                             reported
                                                   600
                                                   800
                                                  1200
                                                  1500
                                                  2500
                                                  4000
                                                  6000

         Guinea-pigs
         (Hartley)           M/6       3-4 weeks     0                 30            5-10           9-20
                                                     1
                                                     5
                                                    10
                                                    15

         Monkeys
         (Macaca             F/2       2.0-3.7 kg    0                 60            1000           14-31
         mulatta)                                  300
                                                  1000
                                                  1500
                                                                                                                             

    a  From: Moore et al. (1979).
    b  M = male;  F = female.
    c  Doses were given orally in corn oil.
    


    10.1.4  Studies on rabbits

         When the above-mentioned soot (or benzene extracts thereof) was
    applied dermally to New Zealand white rabbits (Silkworth et al., 1982)
    (see section 7.4.4.1), it produced no overt toxicity, weight loss, or
    histological changes in thymus, kidney, or skin, but centrilobular
    hypertrophy was found in both sexes. The soot extract gave rise to a
    reversible skin inflammation and hepatic centrilobular hypertrophy in
    females only. Histological examination showed no changes in kidney,
    thymus, and skin.

    10.1.5  Studies on monkeys

         The single oral LD50 value for TCDF in the young female rhesus
    monkey (Macaca mulatta) was found to be 1000 g/kg in a 60-days
    study within the dose range 0, 500, 1000, and 1500 g/kg and with two
    or four animals at each dose level (Moore et al., 1979). The two
    monkeys that received the sublethal dose developed skin lesions and
    had decreased body weight gain. With lethal doses the following overt
    signs occurred after 7 to 10 days: progressive weight loss, loss of
    body fat, facial oedema, loss of facial hair, loss of finger and toe
    nails, and thickening of skin. Death occurred within 2 to 4 weeks.
    Major histological findings included hyperkeratosis of the skin,
    thymic atrophy with lymphoid hypoplasia, and adverse effects on
    epithelial linings. No structural liver lesions were observed though
    the liver weight was increased. Increases in serum albumin and
    cholesterol were also recorded.

         Three male rhesus monkeys given a single dose of 30.6 g TCDF/kg
    body weight did not gain weight during the three following weeks, and
    they developed facial skin lesions, mainly of sebaceous glands
    (Birnbaum et al., 1981).

    10.2  Short-Term Toxicity

    10.2.1  Studies on rats

         Male Sprague Dawley rats were fed 1 or 10 g/kg of a PCDF
    mixture, containing two tetraCDFs, four pentaCDFs, and four hexaCDFs
    for 4 weeks (Oishi et al., 1978). Both diets gave rise to decreases in
    growth rate, food consumption, haemoglobin and haematocrit values,
    erythrocyte counts, serum levels of triglyceride, testosterone,
    glutamic pyruvic transaminase, and leucine aminopeptidase activities,
    as well as increases in serum cholesterol, cholinesterase, and
    glutamic oxaloacetic transaminase activities. Rats fed the 10 g/kg
    diet developed chloracne-like lesions on the ears within 3 weeks.
    Furthermore, this diet decreased the relative weights of thymus,
    prostate, and seminal vesicles and increased the relative weights of
    liver, testes, spleen, adrenals, lung, heart, and brain. In another
    similar study, no effects were seen on total serum proteins or
    leukocyte counts (Oishi, 1977).

         In studies by Hori et al. (1986), male Sprague Dawley rats (aged
    5 weeks) were for 21 days given daily oral doses of a mixture of PCBs,
    PCQs, and PCDFs having a similar composition and isomeric ratio to
    those found in the contaminated rice oil causing "Yusho" (section
    11.1). The toxicities noted included thymic atrophy, suppression of
    weight gain, hepatic enlargement, and an increased serum cholesterol
    level and a decrease in serum glutamic pyruvic transaminase activity.
    The mixture caused an induction of the AAH drug-metabolizing enzyme
    similar to that caused by PCDFs alone. These results support the
    hypothesis that the predominant etiology of "Yusho" involves PCDFs
    contained in the PCB-contaminated toxic rice oil.

    10.2.2 Studies on mice

         Mice (C57Bl/6), given TCDF orally 5 times per week for 30 days
    did not develop clinical signs of toxicity at doses of 30, 100, or 300
    g/kg body weight. However, thymus atrophy, liver hypertrophy,
    decreased leukocyte count, and slightly elevated total serum protein
    did occur in the high dose group at the end of the study (Moore et
    al., 1979). Daily doses of 10, 30, or 50 g TCDF/kg body weight on
    gestation days 10-13 produced a dose-related increase in maternal
    liver weight in C57Bl/6 mice (Weber et al., 1984). Decreased thymus
    weight was recorded in ICR/JCL mice exposed to four weekly doses of a
    mixture (at 100 g/kg) of 12% tetraCDF and 88% pentaCDF (Oishi &
    Hiraga, 1980). When PCDFs of unknown composition were given in the
    diet at 0.6 mg/kg to mice for 10 weeks, severe dermal lesions,
    hyperkeratosis, and dilated hair follicles filled with keratinous
    material occurred in 7 of 12 mice. Furthermore, hepatocytes had
    enlarged nuclei and vacuolations in the cytoplasm (Nagayama et al.,
    1979). Feeding female ddN mice 0.6 mg PCDFs/kg diet (48% tetraCDFs,
    49% pentaCDFs, and 3% hexaCDFs) for 18 days after mating, or for 14
    days after delivery, produced no overt toxic effects in dams or in
    offspring (Nagayama et al., 1980).

    10.2.3 Studies on guinea-pigs

         In studies by Luster et al. (1979b), oral administration of 0.05,
    0.17, 0.5, or 1.0 g 2,3,7,8-TCDF/kg body weight once weekly for 6
    weeks to young female Hartley guinea-pigs produced 30% mortality in
    the high-dose group. The thymus weight was decreased in the 0.5 and
    1.0 g/kg dose groups, though histologically only a slight decrease in
    the density of thymic cortex was observable. Reduction in spleen
    weight or alterations in splenic morphology did not occur, neither was
    there a consistent decrease in body weight.

         Four adult male Hartley guinea-pigs were given six or seven
    weekly doses of 1 g TCDF/kg body weight (Decad et al., 1981a). The
    animals started to lose weight rapidly after the fifth or sixth dose,
    the cumulative dose then being comparable to the oral LD50 value for
    young guinea-pigs. At this time all animals were moribund and by day
    44 the first animal died. Neither hepatomegaly nor thymic atrophy was

    observed in this study. Thus multiple sublethal doses of TCDF appear
    to have a cumulative effect, and may lead to a critical body burden
    that will result in irreversible and progressive weight loss
    eventually followed by death.

         Weekly oral doses of 1 g TCDF/kg body weight or biweekly doses
    of 2 g TCDF/kg body weight (interrupted by a 4-week period of no
    dosing after the fourth dose) to young male Hartley guinea-pigs in
    groups of four resulted in deaths on days 47, 51, 84; 31, 38, 88; and
    32, 70, 85, respectively (Ioannou et al., 1983). At each dosing
    schedule one animal was sacrificed at 101 days after exposure.
    Repeated small doses, with various intervals in between, resulted in
    a similar lethality but a less dramatic weight loss than with a high
    acute dose.

    10.2.4  Studies on rabbits

         The 25% ether-hexane extracts from two commercial polychlorinated
    biphenyl (PCB) preparations containing tetraCDFs and pentaCDFs
    produced hyperplasia and hyperkeratosis of the follicular epithelium
    of the rabbit ear skin when applied dermally weekly for 3 weeks in a
    dose corresponding to 200 mg PCB. Liver lesions or decreased weight
    gain were not observed. No dermal effects could be found when an
    ether-hexane extract from a PCB preparation lacking PCDF impurities
    was applied in the same manner (Vos & Beems, 1971).

         A mixture of tetraCDFs and pentaCDFs was much less potent than
    was TCDD in producing hyperkeratosis when applied to the inside of
    depilated rabbit ears for 3 consecutive days (Nishizumi et al., 1975).

    10.2.5  Studies on hamsters

         No toxic effects were reported in male Golden Syrian hamsters
    (50-70 g) given a diet containing 2.5% HCl-pretreated fly ash from a
    municipal incinerator (Zaanstad, The Netherlands) for up to 95 days
    (van den Berg et al., 1986b).

    10.2.6  Studies on monkeys

         A two-month study with three young male rhesus monkeys (Macaca
    mulatta), serving as their own controls and fed a diet with 50 g
    TCDF/kg, resulted in one case of illness after 1 month and one death
    after 2 months when the cumulative dose was calculated to be 300 g/kg
    (McNulty et al., 1981, 1982a). Toxic changes observed after 1 month
    included periorbital oedema, reddening and thickening of the eyelids,
    enlargement of facial hair follicles, and decreased number and size of
    sebaceous glands in the skin. After 2 months these changes had become
    more severe and were accompanied by decreased physical activity and
    elevated (or eventual loss of) toe and finger nails. There were no
    changes in haematology or serum chemical values. The diseased and the
    surviving monkeys recovered rapidly when they were returned to

    uncontaminated food. Within 3 months, behaviour, clinical appearance,
    and histological structure of the skin were normal. The monkey that
    died had lost 23% of its initial weight and most of its body hair.
    Sebaceous glands were replaced by small squamous cysts. Severe lesions
    were confined to the skin, thymus, and the stomach epithelium, whereas
    liver lesions were modest. Decreased bone marrow cellularity was a
    postmortem finding which was not reflected in the peripheral blood
    count taken before death.

    10.2.7  Studies on chickens

         Mortality in one-day-old White Leghorn chickens given 1 or 5 g
    TCDF/kg body weight orally for 3 weeks was 16% and 100%, respectively,
    with an average time to death of 19 and 11.5 days (McKinney et al.,
    1976). Body weight gain and food consumption were decreased during the
    third week post-treatment. Dose-related subcutaneous oedema, ascites,
    and hydropericardium, as well as thymus atrophy, occurred. Depletion
    of lymphatic cells was evident both in the spleen and thymus. Mild
    liver lesions were found only in the high-dose group. Total serum
    protein and serum albumin were reduced.

         The significant difference in toxicity in chickens between three
    commercial PCB preparations (Vos & Koeman, 1970) was later
    demonstrated to be caused by the presence of tetra- and pentaCDFs in
    two of the three preparations (Vos et al., 1970). The 25% ether-hexane
    extracts from these two PCB preparations were highly toxic in the
    chick embryo assay (Vos et al., 1970), whereas no effect could be
    produced by the extract from the PCB preparation lacking PCDF
    impurities.

    10.3  Chronic Toxicity

    10.3.1  Studies on monkeys

         In studies by McNulty et al. (1981, 1982a), three young male
    rhesus monkeys (Macaca mulatta) were exposed for 6 months to 5 g
    TCDF/kg diet, and one animal served as a control. One animal was
    killed after 6 weeks when moribund. Overt toxic signs in the two
    remaining animals started to appear after 3 months and the symptoms
    remained for the following 3 months. One of these animals died
    suddenly after 6 months. The remaining monkey was returned to normal
    food and rapidly recovered. Clinically and pathologically, chronic
    intake of small amounts of TCDF caused symptoms similar to those
    following a single large dose of TCDF (section 9.1.2.4) or acute or
    chronic ingestion of TCDD (see sections 7.1.1 and 7.3). The major
    histopathological changes in all cases were seen in the thymus,
    sebaceous glands, nail beds, bone marrow, and mucosa of the stomach
    and bile ducts. The toxic potency of TCDF when ingested chronically
    was approximately equal to that of TCDD. This contrasts with the acute
    toxic effect of TCDF, which is approximately 20 times less than that
    of its TCDD counterpart. The reason for death in the TCDF-poisoned

    monkeys was obscure; it was preceded by weight loss, anorexia, and
    depression. Only modest thymic and epithelial changes were present,
    and there was no evidence for liver damage. The quick recovery of
    animals returned to normal diet contrasted with the course of TCDD
    poisoning in which illness progressed to death, or recovery was much
    delayed, even after exposure had ended.

    10.4  Effects Detected by Special Studies

    10.4.1  Immunobiological effects

         To date no studies have been performed on the effects of PCDFs on
    the developing immune system.

         Comparative studies on humoral immune responses in mice have
    revealed that TCDF produces a pattern of responses similar to that
    found for TCDD but only at 30-fold higher doses. Furthermore the
    immunosuppressive effect of TCDD is much more persistent (Vecchi et
    al., 1983b).

    10.4.1.1  Histopathology

         During toxicity studies with pure isomers of PCDFs or with
    mixtures of PCDFs, thymus atrophy has been noted as a consistent
    effect in the mouse (Nishizumi et al., 1978; Moore et al., 1979), rat
    (Oishi, 1977; Oishi et al., 1978), guinea-pig (Moore et al., 1979),
    and monkey (Moore et al., 1979; McNulty et al., 1981). Studies aimed
    at investigating immunobiological effects revealed decreased thymic
    weights (Luster et al., 1979b; Vecchi et al., 1983). The histological
    findings are similar to those occurring after TCDD exposure, i.e.,
    loss of lymphoid cells in the thymic cortex. A reduced number of
    spleen cells was obtained from mice treated with a single ip dose of
    180 g TCDF/kg body weight (Vecchi et al., 1983), but no splenic
    pathology was reported in mice given four weekly oral doses of a PCDF
    mixture (at 100 g/kg) containing 12% tetraCDFs and 88% pentaCDFs
    (Oishi & Hiraga, 1980). Peritoneal cell and macrophage counts in mice
    were not modified by an ip dose of TCDF (180 g/kg body weight)
    (Vecchi et al., 1983).

    10.4.1.2  Humoral-mediated immunity

         Adult female Hartley guinea-pigs exposed orally to 0.05, 0.17,
    and 0.5 g TCDF/kg body weight once weekly for 6 weeks showed somewhat
    depressed serum IgG concentrations. A dose-related depression in
    splenic lymphocyte proliferation was seen in TCDF-treated animals
    after stimulation with the B-lymphocyte mitogen Escherichia coli
    0127 lipopolysaccharide at 50 g/ml medium. There were no effects on
    any of the major serum proteins, neither was there an effect on the
    antibody response towards bovine gamma globulin (BGG) (Luster et al.,
    1979b).

         The antibody response to sheep red blood cells given 7 days after
    a single ip injection of 180 g TCDF/kg body weight was inhibited by
    85% and 35% in C57Bl/6 and DBA/2 mice, respectively (Vecchi et al.,
    1983), whereas a single ip dose of 10 g TCDF/kg body weight to
    C57Bl/6 mice had no effect (Rizzardini et al., 1983). The suppression
    noted by Vecchi et al. (1983) was dose dependent as well as time
    dependent; by day 42 post-treatment a near-normal antibody response
    was obtained.

    10.4.1.3  Cell-mediated immunity

         Oral intubation of 10 or 100 mg PCDF (12% tetraCDFs and 88%
    pentaCDFs) per kg body weight once weekly for four weeks increased the
    sensitivity to endotoxin of ICR/JCL mice. Following an ip injection of
    50, 250, or 500 g endotoxin per mouse a dose-dependent increased
    mortality was noted two days after the final treatment with PCDF
    (Oishi & Hiraga, 1980). Only at high dose levels were there any
    effects on cell-mediated immunity functions in female Hartley
    guinea-pigs given 0.05, 0.17, 0.5, or 1.0 mg TCDF/kg body weight
    orally once weekly for six weeks (Luster et al., 1979b).

         Both the depression in delayed hypersensitivity response to
    purified protein derivative and the ability of BGG-sensitized
    lymphocytes to release the macrophage inhibition factor were related
    to the dose of TCDF. Splenic lymphocytes from TCDF-treated animals,
    stimulated with the T-lymphocyte mitogen phyto-haemagglutinin (PHA),
    showed a decreased proliferation. On the other hand proliferation of
    splenic lymphocytes stimulated with concanavalin A (Con A), another
    T-lymphocyte mitogen, showed no TCDF-related effect. The increased
    proliferative response to Con A and PHA in thymocytes co-cultivated
    with thymus epithelial (TE) cells or cultivated in TE-conditioned
    medium was inhibited if the TE cells were pretreated with TCDF for 48
    h, thus suggesting a direct effect on TE cells (Osborne et al., 1984).

    10.4.2  Enzyme induction

         Studies discussed below show that PCDFs are potent enzyme
    inducers, the enzyme-inducing potencies varying greatly depending on
    the position as well as on the number of chlorine atoms substituted.
    The structure-activity relationships of the PCDFs with regard to
    enzyme induction are similar to those for PCDDs, with TCDF and
    2,3,4,7,8-pentaCDF being the most potent (Tables 56 and 61).

    10.4.2.1  Studies on rats

         Intraperitoneal doses of 2.5 mg TCDF/kg body weight given once
    daily for three days to female CD rats induced 38- and 3-fold
    increases, respectively, in AHH and UDPGT activities 24 h after the
    final dose. The cytochrome P-450 content was doubled but no effect was
    found on the aminopyrine N-demethylase activity (Goldstein et al.,
    1978).

         Increased AHH and EROD activities were found in the hepatic
    microsomal fraction from immature male Wistar rats 5 days after ip
    injection of single doses of TCDF (1.7 mol/kg body weight) or
    2,3,4,7,8-pentaCDF (0.3, 1.5, 3.0 mol/kg body weight) (Keys et al.,
    1985). This study also detected an alteration by TCDF in the hepatic
    metabolism of testosterone in these rats. Yoshihara et al. (1981) gave
    a single ip dose of 1, 5, or 10 mg PCDF/kg body weight of 13
    individual PCDFs to young male Wistar rats five days prior to the
    determination of hepatic enzyme activities. Congeners having at least
    three chlorine atoms in the lateral positions typically showed
    increased AHH and DT-diaphorase activities, while those congeners
    having no more than two chlorine atoms in these positions were not
    inductive. The cytochrome P-448 content was increased by 5 of the 13
    congeners whereas the benz-phetamine-N-demethylase activity was
    depressed by 7 of the 13. The most potent isomers, TCDF and
    2,3,4,7,8-pentaCDF, were effective at a single dose of 1 g/kg body
    weight. The ranking of the potency for enzyme-inducing abilities did
    not coincide with the hepatic distribution of the test substances.
    Hepatic AHH activity in male Wistar rats was significantly enhanced
    only by TCDF and 2,3,4,7,8-pentaCDF among the 15 individual PCDF
    isomers tested, the dose administered intraperitoneally being 5 g
    PCDF/kg body weight (Nagayama et al., 1983).

         Eight of the 15 PCDF isomers tested increased the pulmonary AHH
    activity from 5-fold to 30-fold. In this study no PCDF-related AHH
    induction was present in the kidney, prostate, thymus, or spleen.
    Bandiera et al. (1984b) investigated the effect of three tetraCDFs and
    three pentaCDFs at doses of 500 and 1000 g/kg body weight,
    respectively, on hepatic AHH, aminopyrine N-demethylase,
    4-chlorobiphenyl hydroxylase, and EROD activities in male Wistar rats.
    The most active compounds, TCDF and 2,3,4,7,8-pentaCDF, were potent
    inducers of the cytochrome P-448-dependent monooxygenases. Some
    induction of microsomal AHH, EROD, and 4-chlorobiphenyl hydroxylase
    was observed also for the TCDF and 1,2,4,7,9-pentaCDF.

         The ED50 values for hepatic AHH (Table 61) and 4-chlorobiphenyl
    hydroxylase induction were established for 15 individual PCDFs in
    immature male Wistar rats 14 days after a single ip injection (Mason
    et al., 1985).

         Significant induction of hepatic AHH activity in male Sprague
    Dawley rats was given only by 3 out of 25 individual PCDFs given as
    single oral doses of 40 g/kg body weight (Doyle & Fries, 1986). The
    active congeners were 2,7-diCDF, TCDF, and 2,3,4,7,8-pentaCDF.

         A mixture of PCDFs, reconstituting the approximate composition
    found in the liver of Yusho victims (7.4% tetraCDF, 6.1%
    1,2,4,7,8-pentaCDF, 19.0% 1,2,3,7,8-pentaCDF, 29.4% 2,3,4,7,8-pentaCDF
    and 39.1% 1,2,3,4,7,8-hexaCDF by weight) was given as a single ip
    injection to male Wistar rats 14 days before measuring the induction

    of cytochrome P-448-related enzyme activities (Bandiera et al.,
    1984a). A dose-related enhancement of AHH and EROD activities was
    found within the range 10 to 400 g PCDF mixture/kg body weight.

    10.4.2.2  Studies on mice

         No induction of cytochrome P-448 content, or of ECOD activities,
    was found 12 days after a single ip injection of 10 g TCDF/kg body
    weight to male C57Bl/6J mice (Rizzardini et al., 1983).

         Nagayama et al. (1985) investigated the AHH-inducing potency of
    TCDF, 2,3,6,7-tetraCDF, 1,2,3,6,7-pentaCDF, 1,2,3,7,8-pentaCDF,
    2,3,4,6,7-pentaCDF, 2,3,4,7,8-pentaCDF, 1,2,3,4,6,7-hexaCDF, and
    1,2,3,4,7,8-hexaCDF in two strains of responsive (C57Bl/6 and AKR/Qdj)
    and two strains of non-responsive (DBA/2 and DDD) mice. All congeners
    were given as single ip doses of 30 g/kg body weight in olive oil. No
    single congener induced the AHH activity above the control level in
    the non-responsive mice. Significantly increased AHH activity was
    found in both responsive strains exposed to TCDF, 1,2,3,7,8-pentaCDF,
    and 2,3,4,7,8-pentaCDF. Mice (C57BL/6) treated with 2,3,4,6,7-pentaCDF
    and 1,2,3,4,7,8-hexaCDF also responded with increased AHH activity.

    10.4.2.3  Studies on chickens

         Hepatic AHH activity in chick embryos was inducible by TCDF,
    2,3,4,7,8-pentaCDF, and 1,2,3,7,8-pentaCDF, with ED50 values of
    0.015, 0.014, and 0.071 nmol/egg, respectively (Poland et al., 1976).
    No induction was produced by unchlorinated dibenzofuran, 2,8-diCDF,
    2,4-diCDF, 2,4,8-triCDF or 1,3,6,7-tetraCDF at the doses tested. There
    were no effects on ALA synthetase, p-nitrophenol-UDPGT and
    testosterone-UDPGT activities. However, a modest increase in
    cytochrome P-450 content was present in one-day-old White Leghorn
    chickens 3 weeks after treatment with a single oral dose of 1 g TCDF
    (Goldstein et al., 1976).

    10.4.2.4  Studies on cell cultures

         Exposure of primary hepatocytes isolated from adult male Wistar
    rats to TCDF for 72 h resulted in a 2-fold increase in AHH induction
    at 10-9 mol/litre and half-maximal induction at 3 x 10-10
    mol/litre. However, no AHH induction was observed with 2,7-diCDF in
    the range 10-11 to 10-8 mol/litre in the same system (Jansing &
    Shain, 1985). A 59-fold increase in AHH activity and a 40-fold
    increase in EROD activity were obtained in rat hepatoma H-4-II E cells
    when exposed to 5 x 10-10 mol TCDF/litre for 3 days (Keys et al.,
    1986). In this same study it was demonstrated that TCDF had an
    additive effect, whereas 1,3,6,8-tetraCDF and 2,4,6,8-tetraCDF had
    counteracting effects on TCDD-induced enzyme induction.

         The EC50 values for AHH and EROD induction (Table 61) have been
    established for 35 individual PCDFs in the rat hepatoma H-4-II E cell
    line (Bandiera et al., 1984b; Mason et al., 1985). AHH and EROD
    activities were determined after exposing the cells to optimal doses
    of PCDFs for 5 days. Unchlorinated dibenzofuran, or 2- and
    3-chlorodibenzofuran did not induce these enzyme activities. EC50
    values for all the remaining congeners varied between 10-4 and 1.3 x
    10-10 mol/litre, the most active inducer being 2,3,4,7,8-PCDF.

         Human lymphoblastoid cell lines, derived from the peripheral
    blood of healthy male and female volunteers of various ages, were
    exposed to eight individual PCDF isomers for 48 h (Nagayama et al.,
    1985b). The AHH inducibility was highly variable between individuals
    but less variable between isomers. In this system TCDF was about half
    as potent as 2,3,4,7,8-pentaCDF, 1,2,3,4,6,7-hexaCDF, or
    1,2,3,4,7,8-hexaCDF, which were equally as potent as TCDD in inducing
    AHH.

         A mixture of PCDFs, reconstituted on the basis of PCDF residues
    in the liver samples from Yusho victims (see section 10.4.2.1), had
    EC50 values for induction of AHH and EROD activities of 1.02 x
    10-10 and 3.23 x 10-10 mol/litre, respectively, in the rat
    hepatoma H-4-II E assay (Sawyer & Safe, 1985). The calculated EC50
    values based on the relative isomer content of the mixture were 3.07
    x 10-10 and 4.43 x 10-10 mol/litre, respectively.

    10.4.3  Receptor binding

         The competitive binding of PCDFs to the TCDD receptor protein has
    been studied in vitro both in the hepatic cytosol (Poland et al.,
    1976; Bandiera et al., 1984b) and in the nucleus (Poellinger et al.,
    1982). Poland et al. (1974) investigated the ability of seven PCDF
    congeners to compete with TCDD in binding to the hepatic cytosol
    receptor from C57Bl/6J mice. They found the relative binding
    affinities for TCDF, 2,3,4,7,8-pentaCDF, and 1,2,3,7,8-pentaCDF to be
    37%, 34%, and 38%, respectively, of the binding affinity between TCDD
    and the receptor. The EC50 values for the competitive binding of 33
    individual PCDFs to the receptor from rat hepatoma H-4-II E cell
    cultures varied from less than 10-3 mol/litre for
    4-chlorodibenzofuran to 1.5 x 10-8 mol/litre for the most active
    competitor, 2,3,4,7,8-pentaCDF, which had an EC50 value comparable
    to that for TCDD, i.e., 1.0 x 10-8 mol/litre (Table 61) (Bandiera et
    al., 1984b; Mason et al., 1985). Of the TCDD bound to the nuclear
    receptor in vitro, 58% was displaced by a 100-fold molar excess of
    TCDF. These nuclei were isolated from the liver of Sprague Dawley rats
    pretreated intravenously with 1 g TCDF 2 h prior to the incubation
    (Poellinger et al., 1982).

    10.5  Embryotoxicity and Reproductive Effects

         TCDF has been found to be a potent teratogen in mice at doses
    that produce no overt toxic effects in dams. Malformations observed
    include cleft palate and kidney malformation similar to
    hydronephrosis. Dose-related increases in fetal mortality occur with
    single high doses. The teratogenic pattern of TCDF thus is strikingly
    similar to that of TCDD (see section 7.5).

         Single doses of 100 to 1000 g TCDF/kg body weight to pregnant
    C57Bl/6 mice on gestation days 10 to 13 produced dose-related
    increases in the number of cleft palates and kidney malformations;
    both the number of litters and the number of fetuses were affected
    (Hassoun et al., 1984a; Weber et al., 1984). No other
    treatment-related malformations were reported. A cleft palate
    incidence of 40% was obtained in NMRI mice offspring after sc
    treatment of the dams on gestation days 9 to 11 with 200 nmol TCDF/kg
    body weight (Krowke, 1986).

         Palatal closure in mice occurs late on day 14 of gestation, and
    so it is somewhat peculiar that the peak sensitivity for cleft palate
    occurs on day 12 (Hassoun et al., 1984a). The peak sensitivity for
    kidney malformation in mice occurs on day 11 of gestation (Hassoun et
    al., 1984a). The quantitative data on this malformation somewhat
    conflict in the two studies. Weber et al. (1984) reported that 95.5%
    of the fetuses had kidney malformations after a dose of 500 g/kg body
    weight on day 10 of gestation. However, only 17% of the fetuses per
    dam had this malformation after a single dose of 400 g/kg body weight
    on the same day in the study of Hassoun et al. (1984a). The difference
    might be due to unequal judging of the malformation. Preliminary
    results (Weber et al., 1984), suggested that TCDF-induced kidney
    malformations, up to a certain degree, represent a reversible defect
    since no hydronephrotic kidneys were found in neonates, whereas in
    identically treated dams examined on day 18 of gestation over 80% of
    the fetuses/litter were affected. Fetal mortality increased in a
    dose-related manner with high single doses administered on days 10 to
    12. Peak sensitivity occurred on day 10 (Hassoun et al., 1984a).
    Multiple low dosing on gestation days 10 to 13 was more effective in
    producing fetal malformations, but less effective in producing fetal
    deaths, than single high dosing on day 10 (Weber et al., 1984). No
    effect on fetal mortality (days 12, 13, and 14) was observed in
    C57BL/6N mice given a single oral dose of 800 mg TCDF/kg body weight
    on day 11 of gestation (Weber & Birnbaum, 1985).

         Recombinant inbred strains of C57Bl/6 and DBA/2 mice segregating
    at the Ah locus respond differently to the teratogenic effect of TCDF
    (Hassoun et al., 1984b). Fetuses of Ah-responsive strains responded
    with a high frequency of cleft palates and kidney malformations after
    a single ip dose of 600 g TCDF/kg body weight on day 12 of gestation.

    However, no cleft palates and only modestly increased numbers of
    kidney malformations in a few strains were found with the same
    treatment in Ah-nonresponsive strains.

         A diet containing 0.6 mg PCDFs/kg (48% tetraCDFs, 49% pentaCDFs,
    and 3% hexaCDFs), fed to mice for 18 days after mating, had no effect
    on the number or body weight gain of the offspring, neither were there
    any malformations related to the diet (Nagayama et al., 1980).

         Three PCDFs, namely 1,2,3,7,8-pentaCDF, 2,3,4,7,8-pentaCDF, and
    1,2,3,4,7,8-hexa CDF, are teratogenic to C57BL/6N mice when
    administered orally by gavage on gestation days 10-13. A significant
    increase in hydronephrosis and cleft palate was found, with
    2,3,4,7,8-PCDF being the most potent PCDF studied, having an ED50 of
    36 g/kg body weight for cleft palate and 7 g/kg for hydronephrosis.
    For all three PCDFs, hydronephrosis occurred at a lower dose than did
    cleft palate (Birnbaum et al., 1987).

         It has been pointed out by McNulty (1985) that chlorinated
    compounds such as 2,3,7,8-TCDD produce cystic periodontal lesions and
    squamous metaplasia of the ameloblasts surrounding unerupted teeth in
    rhesus monkeys. These findings are similar to those on the teeth
    development seen in Yusho patients (section 11.1).

    10.6  Mutagenicity

         When tested in Salmonella typhimurium strains KTA98 and TA100
    with and without metabolic activation, no mutagenic activity was found
    for 2,9-diCDF, 3,6-diCDF, TCDF, or octaCDF (Schoeny, 1982). TCDF was
    also studied in Saccharomyces cerevisiae strain MP-1 and was found to
    be negative for forward mutation, mitotic crossing over, and mitotic
    gene conversion at concentrations up to 1000 mg/litre. Stationary
    phase cells were tested in the absence of exogenous activation (Fahrig
    et al., 1978).

    10.7  Carcinogenicity

         The hepatic tumour-promoting activity of a commercial
    polychlorinated biphenyl mixture, Aroclor 1254, with (Ar 1254) or
    without (Ar 1254-PCDF) PCDF impurities, was studied in Sprague Dawley
    rats pretreated with 66 g diethylnitrosamine/ml drinking water for 5
    weeks (Preston et al., 1981). Thereafter the rats were fed a diet
    supplemented with 100 g Ar 1254/kg (> 3 mg PCDF/kg) or Ar
    1254-PCDF (<0.1 mg PCDF/kg) for 18 weeks. Examination of liver
    lesions by light microscopy demonstrated that Ar 1254 promotes
    formation of hepatocellular carcinomas in rats. The promoting
    incidence of 64% remained essentially unchanged when PCDF was removed
    from Ar 1254 by adsorption chromatography. Due to the high incidence
    of hepatocellular carcinomas produced by Ar 1254-PCDF itself, an
    additional effect of PCDF might have been difficult to measure in this
    study.

    11.  EFFECTS OF PCDFs ON HUMAN BEINGS

         Braun (1955) was the first to report chloracne due to chlorinated
    dibenzofurans, subsequently experimentally proven by Bauer et al.
    (1961). Vos et al. (1970) identified by mass spectrometry the presence
    of chlorinated dibenzofurans in commercial PCB mixtures, accounting
    for their acnegenic properties.

    11.1  Yusho and Yu-cheng

         A mass outbreak of food poisoning occurred in western Japan in
    1968 following ingestion of a commercial brand of rice oil
    contaminated with polychlorinated biphenyls (PCBs) and related
    hydrocarbons. The poisoning was named "Yusho" (oil disease).
    Epidemiological proof of the cause of the epidemic depended on the
    demonstration of a dose-response relationship between the consumption
    of the toxic rice oil and the incidence of the poisoning or between
    the oil consumption and the clinical severity of the reaction.
    Approximately 2000 cases were recognized. In 1969, Japanese scientists
    first reported that the toxic rice oil which caused Yusho was
    contaminated with polychlorinated biphenyls (Tsukamoto et al., 1969).
    A few years later, the oil was found also to be contaminated with a
    smaller quantity of PCDFs (Nagayama et al., 1976) and a relatively
    large amount of polychlorinated quaterphenyls (PCQs) (Masuda &
    Yoshimura, 1982).

         In March 1979, an epidemic of a peculiar skin disease broke out
    in Taichung and Changhwa in Central Taiwan. The cause of the disease
    was later identified to be the ingestion of rice-bran oil contaminated
    with polychlorinated biphenyls (Chen et al., 1980, 1981; Hsu et al.,
    1984; Masuda et al., 1986). By the end of 1980, the total number of
    reported cases was about 2000. The local name for the disease was
    Yu-cheng. The mean consumption of total PCDFs of the Yusho and
    Yu-cheng patients has been estimated to be 3.3-3.8 mg/person or
    400-500 mg of toxic 2,3,7,8-substituted PCDFs per person. (Hayabuchi
    et al., 1979). Hayabuchi et al. (1979) estimated the daily intake of
    total PCDFs in the Yusho intoxication to have been 0.9 g/kg body
    weight. Analyses of liver samples taken from the Yusho patients about
    18 months after the exposure showed a dramatic decrease in the number
    of PCDF isomers. Apparently most of the PCDF isomers were metabolized
    or excreted during the period between exposure and sampling (Rappe et
    al., 1979). A comparison between the PCDF isomers found in the Yusho
    oil and the liver samples revealed an interesting relationship. Most
    of the isomers retained had all lateral positions (2-, 3-, 7-, and 8-)
    substituted with chlorine (Rappe et al., 1979).

        Table 72. Clinical symptomatology of Yusho 1969-1972a

                                                                               
    1.   Skin (82-87%).
         Acneiform eruptions, districtive hair follicles, red plaques on limbs,
         dark brown pigmentation of nail, skin, and mucous membranes, itching,
         sweating of palms.

    2.   Ocular manifestations (83-88%).
         Increased eye discharge, swelling of the upper eyelids, hyperaemia of
         conjunctiva, transient visual disturbance.

    3.   Jaundice (10%).
         No abnormalities of liver function in the majority of cases.

    4.   Numbness of the limbs, feeling of weakness, muscular spasms (32-39%).
         Reduced sensory and motor nerve conduction velocity in a few cases
         (9%).

    5.   Hearing difficulties (18%).

    6.   Headaches, vomiting, diarrhoea (17-39%).

    7.   Chronic bronchitis (40%).
         Low serum IgA and IgM, PCB in the sputum.

    8.   Irregular menstrual cycles (60%).

    9.   Dark brown skin pigmentation (which gradually fades) of newborn,
         retarded growth, abnormal teeth number and shape.
                                                                               


    a     Numbers in brackets refer to per cent of patients exhibiting the
         symptoms.
    
        Table 73. Changes in the clinical symptomatology of Yusho
    in the years 1968-1978 
                                                                               
    1.   Skin lesions.
         All skin symptoms diminished gradually, subcutaneous cyst formation
         still present in some of the most severe cases.

    2.   Ocular manifestations.
         Eye discharge, oedema of the eyelids, pigmentation of eyelids and
         conjunctiva, and cyst formation of tarsal gland still present in some
         of the cases.

    Table 73.(cont'd)  Changes in the clinical symptomatology of Yusho
    in the years 1968-1978 

                                                                               

    3.   Stomatological alterations.
         Pigmentation of oral mucosa decreased gradually; anomalies in number
         of teeth and shape of the root still present.

    4.   Chronic bronchitis correlated in severity with concentration of PCBs
         in sputum and blood.

    5.   Serum triglycerides.
         The hyperglycidaemia observed in 1968-1970 returned to a normal level
         by 1973 in females and by 1975 in males.

    6.   Mortality.
         Of 737 cases in the Fukuoka region, 51 (6.92%) died between 1968 and
         1978; there were 11 cancer deaths (3 stomach cancer, 2 lung cancer,
         1 breast cancer, 1 liver cancer, 2 malignant lymphoma).

                                                                               
    

         The rate of excretion of these toxic PCDF isomers is very slow.
    Rappe et al., (1983c) could detect 2,3,4,7,8-pentaCDF in blood plasma
    from Yusho patients when the samples were collected 11 years after
    exposure. Higher levels were found in blood from Yu-cheng patients one
    year after exposure and these analyses also showed a 15-20% reduction
    in one year (Rappe, 1984). PCDFs are selectively retained in the
    liver, with levels corresponding to the fatty level of the tissue.
    They are not found in unexposed controls or in PCB-exposed workers.
    PCB levels in Yusho patients were only about two times higher than
    those of normal people several years after the outbreak. PCB-exposed
    workers had more than 10 times greater PCB blood levels than Yusho
    patients, whereas the PCQ levels of these two groups were similar.

         Generally a correlation between degree or severity of clinical
    signs and the amount of PCDFs retained in the blood exists, whereas
    there is no correspondence between the severity of disease and PCB
    concentrations in blood.

         Mild dermal lesions seen in workers exposed to PCB disappeared
    quickly after discontinuation of PCB handling, in contrast to the
    persistence of Yusho and Yu-cheng symptoms. Everything thus suggests
    that the PCDF contaminant is the causative agent.

         Immunological evaluation of patients exposed in 1979 in Taiwan
    (Yu-cheng) has been reported by Chang et al. (1981, 1982a, 1982b) and
    Wu et al. (1984). Serum immunoglobulin concentrations and lymphocyte
    subpopulations were determined in the peripheral blood of 30 patients

    exposed to PCBs and 23 healthy individuals. The groups were age and
    sex matched. In the patients, serum concentrations of IgA and IgM, but
    not of IgG, were significantly decreased. Also the percentages of
    total T lymphocytes and T-helper lymphocytes were significantly
    reduced, whereas the percentages of T-suppressor cells and B
    lymphocytes were not affected (Chang et al., 1981).

         In a later report (Chang et al., 1982a), monocyte and
    polymorphonuclear lymphocyte (PMN) complement and Fc receptors were
    evaluated in peripheral blood from 30 Yu-cheng patients and 23 normal
    human subjects. Monocytes and PMNs from patients had significantly
    lower percentages of cells bearing immunoglobulin Fc and complement
    receptors. The immune system was further investigated by determining
    the delayed-type hypersensitivity skin response to streptokinase and
    streptodornase, as parameters of cellular immune function. The
    response was studied in 30 PCB-poisoned patients and 50 healthy
    volunteers. Results of the study showed that 80% of the controls had
    positive hypersensitivity skin tests, compared to only 43% of the
    patients. The significant suppression of cellular immunity correlated
    with the severity of the dermal lesions; the size of the
    hypersensitivity skin reaction was negatively correlated with the
    dermal lesions and also with PCB concentrations in whole blood (Chang
    et al., 1982b). More recently, the delayed-type hypersensitivity
    response to tuberculin was reported in 83 PCB-poisoned Yu-cheng
    patients and in 30 age-and sex-matched healthy controls. Compared to
    a positive response rate of 74% in the control group, the patients had
    a significantly lower skin response of 48% (40 out of 83 patients). In
    contrast to the delayed-type hypersensitivity response, the in
    vitro proliferation responses of peripheral blood lymphocytes
    treated with phytohaemagglutinin and pokeweed mitogens, as well as
    tuberculin, were significantly enhanced (Wu et al., 1984).

         There are also indications of immunosuppression in the Yusho
    poisoning. Serum IgA and IgM levels decreased considerably within 2
    years after the onset of the disease. Respiratory involvement included
    bronchiolitis, and respiratory distress was often exacerbated by viral
    or bacterial infection (Shigematsu et al., 1978).

         The symptomatology of Yusho has been summarized by Reggiani
    (1983b) and is to be found in Tables 72 and 73. It is similar to that
    of Yu-cheng, but there are differences such as the frequency of
    transient visual disturbances, hearing difficulties, and a persistent
    bronchitis.

    12.  EVALUATION OF HEALTH RISKS FROM THE EXPOSURE TO CHLORINATED
    DIBENZO-P-DIOXINS (PCDDs) AND DIBENZOFURANS (PCDFs)

    12.1  Introduction

         In order to evaluate the human health risk of PCDDs and PCDFs, it
    is necessary to know both the levels of human exposure and the
    corresponding human health effects.

         Human exposure assessment is complex. Several approaches may be
    taken to estimate it, such as the following.

         (a)  Use of standard physiological models of inhalation,  
              ingestion, and dermal absorption. Data requirements  
              include detailed information on ambient levels in  
              environmental media and food, and on bioavailability.
         (b)  Intake estimates based on simple pharmacokinetic models and
              known levels in human tissue for accurate assessment of
              exposure. Detailed knowledge is also required of the uptake,
              distribution, metabolism, and elimination of PCDDs and PCDFs
              in humans.

    12.2  Exposure Assessment

    12.2.1  Sources of contamination

         The main sources of PCDDs and PCDFs that have so far been
    identified are contaminated commercial chemicals (see section 3.3),
    emissions from combustion sources (see section 3.5), and disposal of
    industrial wastes containing PCDDs and PCDFs (sections 3.4, 3.5.9, and
    3.5.10).

         In some cases estimates of the relative contribution of these
    sources can be generated on a local basis. However, the data and
    methods available today do not allow firm conclusions with regard to
    the relative quantitative importance of these sources on a nation-wide
    or world-wide basis.

    12.2.2  Ambient levels

         The limited data available indicate very low (fg/m3) background
    levels in ambient air of the 2,3,7,8-tetra-, penta-, and
    hexachlorinated PCDDs and PCDFs. (If hepta- and octachlorinated
    congeners are included, pg/m3 levels are noted).

         The few data available indicate that the 2,3,7,8-tetra-, penta-,
    and hexachlorinated PCDDs and PCDFs are unlikely to occur in finished
    drinking-water, even at a level of 1 pg/litre (see section 5.2). In
    all soil and sediment samples analyzed (both from industrialized and
    non-industrialized areas), PCDDs and PCDFs were identified at levels
    ranging from a few ng/kg to several hundred ng/kg (the latter in
    sediments and in urban soil).

         There are no available data on background levels of PCDDs and
    PCDFs in vegetation in the general environment.

         Levels of up to 50 ng/kg of the 2,3,7,8-tetra-, penta-, and
    hexachlorinated PCDDs and PCDFs (principally the tetra- and
    pentachlorinated congeners) have been found in fish from the general
    environment. For the most part these have been detected in fatty or
    bottom-feeding fish (see section 4.4.2).

         Data from terrestrial organisms are inadequate for estimation of
    background levels (see section 4.4.3). Three samples of pooled cow's
    milk showed a maximum of 100 pg/kg of the 2,3,7,8-tetra-, penta-, and
    hexachlorinated PCDDs and PCDFs in whole milk (see section 5.4). Data
    on contamination of other commercial foods are also very limited.
    Analyses of several samples of chicken and pork have shown
    contamination with highly chlorinated congeners at about 5-30 ng/kg.
    The congener profile differs from that noted in aquatic organisms (see
    section 5.4).

         The data available are not sufficient for assessing the total
    exposure of general populations. They are sufficient to perform a
    limited evaluation of exposure for local populations. Based on the
    environmental levels discussed above and the usual assumptions
    regarding intakes of foodstuffs, air, and water, food is more likely
    to be a significant source of PCDD/PCDF exposure than air, while
    drinking-water is likely to be of much less concern.

    12.2.3  Routes of exposure

         Human adipose tissue contains 2,3,7,8-tetra-, penta-, and
    hexachlorinated PCDDs and PCDFs. This contamination is presumably due
    to exposure at the ambient concentrations noted in the preceding
    sections.

         In addition, infants may be exposed through breast milk, and
    small children may also be exposed through ingestion of contaminated
    soil. However, this latter route of exposure, in most instances, is
    likely to be of concern only in heavily contaminated areas.

         Some populations have been at special risk through exposure in
    industrial accidents (and their clean-up) that have occurred during
    the normal production and use of chlorphenols and phenoxy herbicides
    and PCBs. In these situations inhalation and dermal contact are the
    exposure routes of greatest concern. However, quantitative information
    on the nature and concentration of contaminants is not available.

         Based on the environmental levels discussed above and the usual
    assumptions regarding physiological and intake parameters, ingestion
    is likely to be the exposure route of greatest concern. Inhalation of

    ambient air is not likely to be a problem, although inhalation of
    heavily contaminated air may make a significant contribution to
    exposure. In general, it is not possible, at present, to estimate the
    relative contribution of dermal exposure.

    12.2.4  Bioavailability

         No bioavailability data from studies in humans are available.
    From animal studies, it is clear that bioavailability of PCDDs and
    PCDFs following ingestion depends on the matrix ingested. Table 46
    summarizes the data available on oral intake.

         Studies on hairless rats indicate that dermal exposure through
    intact skin from contact with contaminated soil is about 1 to 2%. No
    data are available for inhalation exposure.

    12.3  Animal Data

    12.3.1 Toxicokinetics of 2,3,7,8-TCDD

         Studies on rodents given single or repeated oral doses of
    2,3,7,8-TCDD have shown that 50% or more of the administered amount is
    absorbed from the gastrointestinal tract in rats, guinea-pigs, and
    hamsters, but less than 30% in mice (Table 40). The reported
    half-lives for elimination were in the ranges 12-31 days for rats,
    mice, and hamsters and 22-94 days for guinea-pigs (Table 41). However,
    most of these studies have been performed at toxic doses. The
    half-life of 2,3,7,8-TCDD in primates has not been well established,
    but available data for the rhesus monkey suggest an apparent half-life
    in the adipose tissue of about 1 year.

         2,3,7,8-TCDD does accumulate in animal tissues. In rodents
    accumulation occurs predominantly in the liver and adipose tissue
    (Tables 42 to 44). In rhesus monkeys (Table 45), high levels of
    2,3,7,8-TCDD are recovered from the adipose tissues, liver, skin, and
    muscles.

         At a daily dose of 1 ng 2,3,7,8-TCDD/kg body weight for 2 years,
    rats accumulated 540 ng 2,3,7,8-TCDD/kg body weight in the liver where
    some morphological changes were also observed. Similar levels were
    found in beach mice that had been exposed to 2,3,7,8-TCDD soil levels
    ranging from 10-710 ng/kg. Total body exposure of animals to
    2,3,7,8-TCDD in soil at such concentrations may thus result in tissue
    levels that have been demonstrated to cause effects in experimental
    animals.

         TCDD is largely eliminated in the faeces, although some urinary
    excretion occurs. The hamster has a higher urinary elimination than
    other species studied.

         Transformation of 2,3,7,8-TCDD to more polar metabolites occurs
    in all animal species investigated (see section 6.2 and Table 62).
    Elimination of metabolites from tissues into faeces and urine occurs
    rapidly in all of these species except in the case of the guinea-pig.
    Known metabolites are much less toxic than the parent compound.

    12.3.2  Toxicokinetics of PCDDs and PCDFs, other than TCDD

         Animal data on the toxicokinetics of pure PCDDs other than
    2,3,7,8-TCDD are limited. PCDFs have been more extensively studied in
    this respect. The half-life for 2,3,7,8-TCDF has been reported to be
    in the range of 2-8 days for rats, mice, and rhesus monkeys and more
    than 20 days for guinea-pigs (Table 65). Studies on rats have shown
    that 2,3,4,7,8-pentaCDF is more highly retained than is 2,3,7,8-TCDF
    (65% and 3.8%, respectively, after 5 days).

         Tissue retention data of PCDDs and PCDFs in various species
    exposed to synthetic mixtures or to environmental samples containing
    PCDDs and PCDFs show a high variability in retention time between
    congeners with or without chlorine substitution in all the positions
    2,3,7, and 8.

    12.3.3  Toxic effects of 2,3,7,8-TCDD

         The toxic and biological effects resulting from exposure to
    2,3,7,8-TCDD are dependent on a number of factors, including the
    species, strain, age, and sex of the animals used. The toxic responses
    observed in several animal species include body weight loss,
    hepatotoxicity, porphyria, dermal toxicity, gastric lesions, thymus
    atrophy and immunotoxicity, teratogenicity, reproductive effects, and
    carcinogenicity. TCDD induces a wide spectrum of biological effects
    including enzyme induction and vitamin A depletion. The complete
    spectrum of toxic and biological effects is not usually observed in
    any single animal species. The two most characteristic toxic effects
    observed in all laboratory animals are body weight loss and thymus
    atrophy and immunotoxicity. Chloracne and related dermal lesions are
    the most frequently noted signs of 2,3,7,8-TCDD toxicosis in humans;
    dermal lesions are also observed in rhesus monkeys, hairless mice, and
    rabbits. In contrast, rats, most strains of mice, guinea-pigs, and
    hamsters do not develop chloracne and related dermal toxic lesions
    after exposure to 2,3,7,8-TCDD. Many of the observed toxic lesions are
    either hyperplastic/metaplastic or hypoplastic, and primarily affect
    epithelial tissues.

         Reproductive toxicity has been reported in rhesus monkeys: the
    lowest-observed-effect level (LOEL) was calculated to be 1 to 2 ng/kg
    body weight per day. A no-observed-effect level (NOEL), or possibly a
    LOEL, of 1 ng/kg body weight per day for reproductive effects in rats
    has been discussed (Murray et al., 1979; Nisbet & Paxton, 1982).

         If the cancer studies in rats conducted by Kociba et al. (1978)
    and by the NIH (1982a,b) are compared, it is evident that the liver
    tumours, including hepatocellular carcinomas, are produced at similar
    dose levels. Although an increased incidence of tumours in other
    organs was observed by the NTP, and by Kociba et al. (1978), the other
    target organs varied in the two studies. This may be caused, in part,
    by differences in dosing (gavage versus exposure in ground feed) and
    by differences in strains. In the Kociba study, doses of 10 ng/kg body
    weight caused an increased incidence of neoplastic (hyperplastic)
    nodules in females, and doses of 1 ng/kg body weight resulted in foci
    or areas of hepatocellular alteration (swollen hepatocytes). At these
    dose rates in experimental groups, the incidence of certain
    hormone-dependent tumours was lower than in the control animals,
    suggesting endocrine changes induced by 2,3,7,8-TCDD. Based on these
    animal studies and on available human data IARC (1982 suppl. 4)
    concluded that TCDD showed sufficient evidence for carcinogenicity in
    animals, but inadequate evidence for carcinogenicity in humans.

         TCDD does not appear to have mutagenic properties, and is,
    therefore, not likely to be genotoxic. Thus, it is assumed to be
    carcinogenic through an indirect (epigenetic) mechanism.

    12.3.4  Toxic effects of PCDDs and PCDFs, other than TCDD

         Several other PCDDs and PCDFs cause signs and symptoms similar to
    those of 2,3,7,8-TCDD, but there is a wide variation with regard to
    potency (Tables 56, 62). In summary, there are 12 isomers that display
    high toxicity, i.e., the tetra-, penta-, hexa-, and heptaCDDs and CDFs
    with four chlorine atoms in the symmetrical lateral positions 2,3,7,
    and 8. A mixture of two hexaCDDs (1,2,3,7,8,9- and
    1,2,3,6,7,8-hexaCDD) has been demonstrated to possess carcinogenic
    properties in long-term animal studies, but at higher doses than those
    used in the study of TCDD. Unsubstituted dioxin and 2,7-diCDD failed
    to demonstrate carcinogenic properties.

         The relative toxic and biological potencies of PCDDs and PCDFs
    have been estimated using short-term studies in rats and mammalian
    cell cultures. Endpoints used include inhibition of body weight gain,
    thymic atrophy, enzyme induction, teratogenicity, acnegenic response,
    and keratinization. In the absence of long-term toxicity data, results
    obtained from such short-term tests are at present the only source for
    ranking the toxicity for human risk assessment.

         When investigated, mixtures of these compounds have shown
    additive or less than additive responses.

    12.3.5  Review of species differences

         There are marked species differences in the susceptibility to the
    biological and toxic effects elicited by 2,3,7,8-TCDD. For example,
    the oral LD50 values range from 0.6 g/kg body weight in

    guinea-pigs, to 5051 g/kg body weight in Golden Syrian hamsters
    (Table 47); moreover, pronounced differences in LD50 values have
    also been reported in different strains of the same species (e.g.,
    rats and mice). The toxicity and toxicokinetics of TCDD in monkeys
    most closely resemble the effects observed in human beings. However,
    the tremendous variation in species and strain sensitivity to
    2,3,7,8-TCDD and related compounds cannot be explained by the observed
    toxicokinetic differences. There is evidence in inbred mice, that the
    cellular levels of the Ah receptor correlate, in part, with
    susceptibility to the biological and toxic effects of these compounds.
    The receptor has also been identified in other species, including
    human beings. However, interspecies comparison of cellular Ah receptor
    levels do not explain their differences in sensitivity to
    2,3,7,8-TCDD; this is consistent with complex as yet unknown
    mechanisms of toxicity that involve multiple factors in addition to
    the Ah receptor.

    12.4  Human Health Effects

    12.4.1  PCDDs

         Exposure of the general population is to small amounts of PCDDs
    and PCDFs in complex mixtures and these have not been associated with
    disease. In a few incidents workers and others have been exposed to
    larger amounts of a limited number of these compounds, e.g., Seveso
    and in Yusho disease.

         For occupational and accidental exposure the most prominent
    clinical effect has been chloracne. Other effects (Table 64) have been
    noted, but, apart from chloracne and perhaps minor functional
    disorders, none has been persistent.

         In some, but not all mortality studies, an increased incidence of
    cancer at different sites has been claimed, but the small numbers of
    cases limit confidence in the findings.

         The overall impression from the follow-up studies is that even
    severe acute systemic effects of TCDD are usually reversible, except
    for chloracne, or markedly improved over time following cessation of
    exposure. In Seveso, the only clear-cut adverse health effect recorded
    has been chloracne. 193 cases of chloracne occurred in 1976 and 1977,
    and 20 of those still presented active symptoms in 1984. Many studies
    have been performed to find possible links between exposure and health
    effects in civilians or military personnel exposed to Agent Orange in
    Viet Nam. However, the information available to date does not allow
    definite conclusions to be drawn with regard to effects on human
    reproduction or any other significant health effects (see section
    9.2).

         In a number of studies, exposed populations and various control
    groups have been compared by measuring serum lipids, liver function
    tests, and other variables. Although certain statistically significant
    differences have been reported there, and also in isolated case
    reports, lack of uniformity, various technical shortcomings, and the
    inability to exclude confounding factors means that the results have
    been inconclusive.

         In the Missouri (USA) incident, children who showed acute illness
    when the contamination occurred in 1971 are now reportedly in good
    health. Epidemiological studies in Missouri on populations exposed to
    lower concentration over longer periods of time have so far not
    revealed any significant health effects. Although no clinical illness
    was observed, there were indications of an effect on the cell-mediated
    immune system.

         The ranges of health effects produced by TCDD in human beings
    have yet to be defined. It can be concluded that the data from human
    exposure and effects, when taken together, do not allow any
    determinations of dose/effect or dose/response relationships in human
    beings.

         In spite of many clinical and follow-up studies, no clearcut
    persistent systemic effects have been delineated, except for
    chloracne. In the light of present information, it seems unlikely that
    permanent, severe, and debilitating toxicological sequelae are
    inevitable after exposure to TCDD.

    12.4.2  PCDFs

         The only well documented intoxications with PCDFs in human beings
    are the two instances of contamination of rice oil with PCDFs, PCBs,
    and PCQs, i.e., Yusho in Japan (1968) and Yu-cheng in Taiwan (1979)
    (see section 11.1). In total, several thousand people were acutely
    intoxicated. The summarized data makes it most likely that the
    causative agent was PCDFs. The general symptomatology was similar to
    that found in intoxications with TCDD. The differences may reflect
    intensity in exposure and the ages and sex of the exposed human
    beings. Attempts to estimate the average daily intake of PCDFs over
    several months in Yusho patients indicated a figure of 0.9 g/kg body
    weight of total PCDFs, 0.1-0.2 g/kg of 2,3,7,8-substituted tetra-,
    penta-, and hexaPCDFs, together with 157 g PCBs and 148 g PCQs/kg
    body weight (Hayabuchi et al., 1979). The lowest dose causing disease
    was estimated to be 0.6 mg total PCDFs per person over 30 days,
    corresponding to a daily dose of 0.05-0.1 g/kg body weight of
    2,3,7,8-substituted PCDFs. However, the data available are not
    sufficient to permit any conclusions as to what dose might be safe for
    human intake.

    12.4.3  Human body burden and kinetics

         In human fat, background levels of TCDD up to 20 ng/kg have been
    found in the general population with no known specific exposure, but
    higher levels have been reported in some cases without evidence of
    disease. None of these populations have been randomly sampled. The
    more highly chlorinated other PCDDs and PCDFs, especially octaDD, also
    occur in these samples (see Tables 29, 30). Averages values seem to
    increase with age.

         In special situations, higher levels (in the low g/kg range)
    have been found that have not been associated with disease.

         In the Yusho and Yu-cheng incidents, symptoms were noticed at
    higher levels of PCDFs, e.g., 2,3,4,7,8-pentaCDF was found at 6.9
    g/kg fat tissue one year after the exposure to contaminated rice oil.

         Based on the very limited data available, the levels of, for
    instance, 2,3,4,7,8-pentaCDF in the general population, seem to be two
    orders of magnitude lower than the levels associated with the Yusho
    disease.

         No such comparisons can be made for PCDDs.

         Limited data indicate that those isomers chlorinated at the
    2,3,7, and 8 positions are selectively retained, except for TCDF.

         A half-life for TCDD in human beings of 5 years has been
    indicated by one experimental study. In another study half-lives in
    the range of 2-6 years were estimated for 1,2,3,6,7, 8-hexaCDD,
    1,2,3,4,6,7,8-heptaCDD, octaCDD, 1,2,3,4,6,7,8-heptaCDF, and octaCDF.
    These data need to be expanded since they are based on studies in only
    two subjects and since the toxicokinetics of these types of compounds
    may not simply be controlled by first-order kinetics. However, even if
    there are limitations in the present data, it is apparent that the
    half-lives of these compounds are in the range of one or more years.

         These reported half-lives for human beings are very different
    from those reported in rodents. However, animal experiments have
    usually been performed with toxic doses. Furthermore, animals with a
    short life span have a higher metabolic rate, thus shorter half-lives
    could be expected.

         The PCDDs and PCDFs are predominantly stored in fat, but they are
    also excreted in milk (Table 39) and pass the placenta. They also
    appear in the blood and vital organs at lower concentrations. The
    distribution between different tissues in human beings is not at
    present clear, although it has been suggested that the ratio between
    fatty tissue and liver is higher in human beings than that in rodents.

    However, this conclusion is based on very limited data from autopsy
    specimens. Whether this is relevant for the general population remains
    to be seen.

         The intake route for human beings is at present not very well
    delineated, but it has been assumed that intake from food is the main
    route. However, the human infant represents a special case; because of
    transplacental transfer of these compounds, the neonate might be
    expected to be exposed in utero. Levels measured so far in human
    milk suggest that this food might be an important source of these
    compounds.

         No data are available regarding what dose of PCDDs is toxic to
    human beings. However, in the Yusho and Yu-cheng episodes, total
    intakes of total PCDFs in the range 3.3-3.8 mg/person and total
    intakes of 2,3,7,8-substituted PCDFs in the range 400-500 g/person
    were associated with the disease. No good data are available as to
    what intakes occurred without causing disease.

    12.5  General Conclusions

         PCDDs and PCDFs occur throughout the environment and we all
    probably carry a body burden of them. They have sometimes produced
    complex toxic effects following occupational and accidental exposure.

         Based on the Yusho disease and experiments in sensitive species
    of monkeys, and making assumptions about the relative potencies of
    PCDDs and PCDFs, human beings and certain monkey species may have
    comparable sensitivity to these compounds. However, the uncertainties
    related to the real dose received by human beings and the difficulties
    of assessing toxic effects other than chloracne in our species prevent
    a firm conclusion as to the relative resistance of human beings to the
    toxic effects of these compounds. Exposure should be reduced to levels
    as low as are reasonably practicable.

    13.  RECOMMENDATIONS

    1.   Analytical interlaboratory validation and round-robin studies
         using standardized quality assurance and quality control
         procedures are needed to improve analytical methodology.

         Sampling strategy and analytical procedures and data
         interpretation should be optimized and standardized before
         undertaking surveys.

    2.   Further information is required about the origins and
         environmental distribution and fate of PCDDs and PCDFs. Further
         monitoring data, including time trends and determinations of
         isomer patterns, are required for environmental levels of PCDDs
         and PCDFs, especially for food, ambient air, and sediments.

    3.   Data should be obtained about the effects of PCDDs and PCDFs on
         environmental biota.

    4.   More information is required about the bioavailability of PCDDs
         and PCDFs from different matrices in the environment and from the
         diet. Exposure from these sources should be correlated with
         agricultural and industrial practices.

    5.   Simpler and less expensive methods suitable for screening should
         be developed and validated.

    6.   Studies to determine the mechanisms of toxicity of PCDDs and
         PCDFs are needed to support an evaluation of the differences in
         effects between species and to allow extrapolation to human
         beings.

    7.   Further investigation of immunotoxicity is important, including
         cytotoxic T lymphocyte function. Studies of the effects of
         perinatal exposure and of the duration of actions on the immune
         system are important.

    8.   Long-term toxicity studies, including multigeneration
         reproductive studies, in different species with three of the most
         widespread PCDDs and PCDFs, namely 2,3,4,7,8-pentaCDF,
         1,2,3,7,8-pentaCDD, and octaCDD, should be carried out.

    9.   Because humans are exposed to complex mixtures of PCDDs and
         PCDFs, test systems, including techniques applicable to
         evaluating human tissues, should be further developed and
         validated for the toxic potency of these compounds and other
         mixtures. These systems can be used to study mechanisms of
         action, structure activity relationships, and interactive
         effects.

    10.  Investigations to examine the body burden and to correlate it
         with clinical effects and laboratory findings are indicated.
         Follow-up studies of previously exposed groups are important.

    14.  EVALUATIONS BY INTERNATIONAL BODIES AND THE CONCEPT OF TCDD
    EQUIVALENTS

    14.1  International Evaluations

         IARC evaluated the carcinogenic risk of TCDD to man (IARC, 1977,
    1982) and concluded that there was sufficient evidence that it was
    carcinogenic to animals, but that the data for carcinogenicity to
    human beings was inadequate. None of the other PCDDs or PCDFs have
    been evaluated by IARC.

         Regulatory standards for TCDD and mixtures containing TCDD,
    established by national bodies in different countries and the European
    Economic Community, are summarized in the Legal File of the
    International Register of Potentially Toxic Chemicals (IRPTC, 1987).

    14.2  Methodologies Used in Assessment of Risk from PCDDs and PCDFs

    14.2.1  Individual Congeners

         As shown in this monograph, sufficient high quality data for
    assessing human health risks exist only for TCDD. For the other
    congeners and isomers where data do exist, they are generally derived
    from studies using acute exposures in experimental animals and/or from
    in vitro tests.

         The several risk evaluations on TCDD from various countries have
    utilized the long-term ocogenic rat studies (Kociba et al., 1978) or
    the reproduction studies on rats (Murray et al., 1979) or monkeys
    (Schantz et al., 1979). Mathematical models have been applied to the
    cancer data and "virtually safe doses" between 0.006 and 0.028 pg/kg
    body weight per day have been calculated (Kimbrough, 1984). The
    biological relevance of such models has been questioned, since TCDD
    has not been shown to be genotoxic, and has been found to be a strong
    promoter of liver tumours in a two-stage precarcinogenesis study
    (Pitot et al., 1980) To avoid the use of mathematical models, several
    evaluations have used safety factors in the range of 100 to 1000
    applied to the assumed no-effect, or lowest-observed-effect levels in
    the cancer study of Kociba et al. (1978) or the reproduction studies
    of Murray et al. (1979) or Schantz et al. (1978). Using this
    methodology "tolerable daily intakes" have been calculated for human
    beings in the range of 1-10 pg/kg body weight (Denmark, 1984; Ontario,
    1985; US EPA, 1985; Ahlborg & Victorin, 1987).

    14.2.2  Mixtures of PCDD and PCDF congeners and isomers - 
    concept of TCDD toxic equivalents.

         The results of recent isomer-specific analyses of such diverse
    environmental samples as emissions from the combustion of hazardous
    industrial and municipal wastes, soil, industrial process wastes,
    human adipose tissue, and milk indicate that the majority of the 75
    CDD and 135 CDF isomers can be detected. Humans are therefore exposed

    to complex mixtures of these environmental contaminants (sometimes
    2,3,7,8-tetraCDD is only a minor component), and the level of risk
    from such exposures must be assessed.

         In the absence of long-term whole animal tests on complex
    mixtures of PCDDs and PCDFs, as well as similar studies on individual
    isomers and/or congeners, several models have been proposed to relate
    the toxicity of environmental mixtures to the well studied isomer
    2,3,7,8-TCDD. The results from these models are presented as "TCDD
    Toxic Equivalents". A summary of some models which have been published
    is given in Table 74. The toxic potencies used, relative to TCDD, are
    shown. The scientific basis for deriving these relative toxicities is
    somewhat different for each model. The Swiss (Switzerland, 1982) and
    Danish (Denmark, 1984) models are based essentially on the relative
    potency for AHH induction, whereas the German (Germany, 1985) and
    Canadian models (Ontario, 1985) are based on a weighting of all
    available quantitative data. The USA model (US EPA, 1987) utilizes
    primarily the relative carcinogenic potencies of TCDD and hexaCDD,
    with consideration given to other relevant quantitative data. A
    discussion of the structure-activity relationships and relative
    biological activities of many PCDD and PCDF isomers is given in
    sections 7 and 10.

         Data requirements, and assumptions made in the application of
    these models to various environmental mixtures have been reviewed by
    US EPA (1987) and Ontario (1985). Suter-Hofmann & Schlatter (1985) fed
    toluene extracts of acid-washed particulates from a municipal waste
    incinerator to Sprague Dawley rats at levels in the diet corresponding
    to 4, 12, and 24% of particulates. The calculated daily TCDD intakes
    at these doses were 4.8, 14.4, and 28.8 ng/kg body weight. Depending
    upon which model was used from Table 74, the estimated intake of PCDDs
    and PCDFs would be 48-300 (4% particulates), 144-900 (12%
    particulates), and 288-1800 ng TCDD-equivalents/kg body weight per day
    (24% particulates). No mortality was noted in the groups fed 24%
    particulates, but at this dose body weight gain was depressed in
    females and depressed thymus and increased liver weights were noted in
    both sexes. No adverse affects were seen at the 4% and 12% dose
    levels. From the known toxicity of TCDD, more severe effects would
    have been expected from the calculated dose of between 288 and 1800 ng
    TCDD-equivalents/kg body weight per day. These data support the
    conclusion that the application of TCDD-equivalence models may
    over-estimate the inherent risk from exposures to dioxin-containing
    environmental mixtures, depending largely upon the assumptions used in
    deriving the model. However, a definitive conclusion on this point
    awaits further research (US EPA, 1987) (see section 13).



    
    Table 74.  Examples of TCDD-equivalence models
                                                                                                                            
                   Relative PCDD and PCDF toxicities in model

                           Olie et al.    Switzerland   Germany    Denmark     Ontario     United States
    Compound                 (1983)         (1982)      (1985)     (1984)      (1985)      (US EPA 1987)
                                                                                                                            

    MonoCDD                   0                                                0.001          0
    DiCDD                     0                                                0.001          0
    TriCDD                    0                                                0.01           0
    TetraCDD-2,3,7,8          1.0            1.0         1.0        1.0        1.0            1.0
            -non 2,3,7,8      1.0            0.01        0.01                  0.01           0.01
    PentaCDD-1,2,3,7,8        0.1                                   0.01
            -all 2,3,7,8      0.1            0.1         0.1                   0.1            0.5
            -non 2,3,7,8      0.1            0.1         0.01                  0.1            0.005
    HexaCDD-1,2,3,4,7,8                                             0.1
           -1,2,3,6,7,8                                             0.01
           -1,2,3,7,8,9                                             0.01
           -all 2,3,7,8       0.1            0.1         0.1                   0.1            0.04
           -non 2,3,7,8       0.1            0.1         0.01                  0.1            0.0004
    HeptaCDD-1,2,3,4,6,7,8                                          0.01
            -all 2,3,7,8      0.1            0.1         0.01                  0.01           0.001
            -non 2,3,7,8      0.1            0.1         0.001                 0.01           0.00001
    OctaCDD                   0                          0.001                 0.0001         0
    MonoCDF                                                                    0.0001
    DiCDF                                                                      0.0001
    TriCDF                                                                     0.01
                                                                                                                            

    Table 74 (contd).
                                                                                                                            
                                               Relative PCDD and PCDF toxicities in model

                           Olie et al.    Switzerland   Germany    Denmark     Ontario     United States
    Compound                 (1983)         (1982)      (1985)     (1984)      (1985)      (US EPA 1987)
                                                                                                                            
    TetraCDF-2,3,7,8          0.1            0.1         0.1        0.1        0.5            0.1
            -non 2,3,7,8      0.1            0.1         0.01                  0.5            0.001
    PentaCDF-1,2,3,7,8                                              0.2
            -2,3,4,7,8                                              0.2
            -all 2,3,7,8      0.1            0.1         0.1                   0.5            0.1
            -non 2,3,7,8      0.1            0.1         0.01                  0.5            0.001
    HexaCDF-1,2,3,4,7,8                                             0.2
           -1,2,3,6,7,8                                             0.05
           -1,2,3,7,8,9
           -2,3,4,6,7,8                                             0.1
           -all 2,3,7,8       0.1            0.1         0.1                   0.1            0.01
           -non 2,3,7,8       0.1            0.1         0.01                  0.1            0.0001
    HeptaCDF-1,2,3,4,6,7,8                                          0.01
            -1,2,3,4,7,8,9
            -all 2,3,7,8      0.1            0.1         0.01                  0.01           0.00001
            -non 2,3,7,8      0.1                        0.001                 0.01           0.00001
    OctaCDF                   0                          0.001                 0.0001         0
                                                                                                                            
    


         It must be recognized that an approach such as the use of
    "TCDD-equivalents" must be regarded as an interim procedure for the
    measurement of the toxicity of environmental samples in the absence of
    long-term toxicity data on specific PCDD and PCDF isomers and mixtures
    of these compounds. At present it is an imprecise evaluation
    methodology with many data gaps in the supporting data base.

         The TCDD-equivalent models have two major sources of uncertainty,
    i.e., firstly, the unanswered scientific questions related to the
    toxicity of TCDD itself, and secondly, the lack of data on the other
    PCDD and PCDF congeners and isomers that would permit a more accurate
    determination of the potency of these chemicals relative to TCDD. The
    research recommendations in this document address some of these
    concerns. As such data become available, TCDD-equivalent models must
    be continuously updated and risk assessments based on the present
    models (Table 74) considered only as interim evaluations.

    REFERENCES
                                                                       

    ABATE, L., BASSO, P., BELLONI, A., BISANTI, L., BORGNA, C., BRUZZI,
    P., DORIGOTTI, G., FALLIVA, L., FANUZZI, A., FORMIGARO, M., MAGGIORE,
    G., MARNI, E., MEAZZA, L., MERLO, F., PUNTONI, R., ROSA, A., STAGNARO,
    E., & VERCELLI, M. (1982) Mortality and birth defects from 1976 to
    1979 in the population living in the TCDD polluted area of Seveso. In:
    Hutzinger, O., ed. Chlorinated dioxins and related compounds.
    Impact on the environment, Oxford, New York, Pergamon Press, pp.
    571-598.

    ABERNETHY, D.J., GREENLEE, W.F., HUBAND, J.C., & BOREIKO, C.J. (1985)
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) promotes the transformation
    of C3H/10T1/2 cells. Carcinogenesis, 6: 651-653.

    ADAMS, E.M., IRISH, D.D., SPENCER, H.C., & ROWE, V.K. (1941) The
    response of rabbit skin to compounds reported to have caused acneform
    dermatits. Ind. Med., 2: 1-4.

    AHLBORG, U.G. & VICTORIN, K. (1987) Impact on health and environment
    from trace organic emissions. Waste Manage. Res., 5: 203-224.

    AHLBORG, U.G., WAERN, F., & HAKANSSON, H. (1987) Interactive effects
    of PCDDs and PCDFs occurring in human mother's milk. Chemosphere,
    16(8/9): 1701-1706.

    AHLING, B., BJORSETH, A., & LUNE, G. (1978) Formation of chlorinated
    hydro- carbons during combustion of poly(vinyl chloride).
    Chemosphere, 8(10): 799-806.

    AHLING, B., LINDSKOG, A., JANSSON, B., & SUNDSTROM, G. (1977)
    Formation of polychlorinated dibenzo-p-dioxins and dibenzo-furans
    during combustion of a 2,4,5-T formulation. Chemosphere, 6(8):
    461-468.

    AITIO, A. & PARKKI, M.G. (1978) Organ specific induction of drug
    metabolizing enzymes by 2,3,7,8-tetrachlorodibenzo-p-dioxin in the
    rat. Toxicol. appl. Pharmacol., 44: 107-114.

    AITIO, A., PARKKI, M.G., & MARNIEMI, J. (1979) Different effect of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on glucuronide conjugation of
    various aglycones. Studies in Wistar and Gunn rats. Toxicol. appl.
    Pharmacol., 47: 55-60.

    AKERMARK, B. (1978) Photochemical reactions of phenoxy acids and
    dioxins. Chlorinated phenoxy acids and their dioxins. Ecol. Bull.,
    21: 75-81.

    ALBRO, P.W. & CORBETT, B.J. (1977) Extraction and clean-up of animal
    tissues for subsequent determination of mixtures of chlorinated
    dibenzo-p-dioxins and dibenzofurans. Chemosphere, 7: 381-385.

    ALBRO, P.W., CORBETT, J.T., HARRIS, M., & LAWSON, L.D. (1978) Effects
    of 2,3,7,8-tetrachlorodibenzo-p-dioxin on lipid profiles in tissue of
    the Fisher rat. Chem.-biol. Interact., 23: 315-330.

    ALBRO, P.W., CRUMMETT, W.B., DUPUY., A.E., Jr, GROSS, M.L., HANSON,
    M., HARLESS, R.L., HILEMAN, F.D., HILKER, D., JASON, C., JOHNSON,
    J.L., LAMPARSKI, L.L., LAU, B.P.Y., MCDANIEL, D.D., MEEHAN, J.L.,
    NESTRICK, T.J., NYGREN, M., O'KEEFE, P., PETERS, T.L., RAPPE, C.,
    RYAN, J.J., SMITH, L.M., STALLING, D.L., WEERASINGHE, N.C.A., &
    WENDLING, J.M. (1985) Methods for the quantitative determination of
    multiple, specific poly-chlorinated dibenzo-p-dioxin and dibenzofuran
    isomers in human adipose tissue in the parts-per-trillion range. An
    inter-laboratory study. Anal. Chem., 57: 2717-2725.

    ALBRO, P.W., CORBETT, J.T., & SCHROEDER, J.L. (1986) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on lipid peroxidation in
    microsomal systems in vitro. Chem.-biol. Interact., 57: 301-313.

    ALLEN, J.R. (1964) The role of "toxic fat" in the production of
    hydro-pericardium and ascites in chickens. Am. J. vet. Res., 25:
    1210-1219.

    ALLEN, J.R. & CARSTENS, L.A. (1967) Light and electron microscopic
    observations in Macaca mulatta monkeys fed toxic fat. Am. J. vet.
    Res., 28: 1513-1526.

    ALLEN, J.R. & LALICH, J.J. (1962) Response of chickens to prolonged
    feeding of crude "toxic fat". Proc. Soc. Exp. Biol. Med., 109:
    48-51.

    ALLEN, J.R., VAN MILLER, J.P., & NORBACK, D.H. (1975) Tissue
    distribution, excretion and biological effects of (14C)
    tetrachlorodibenzo-p-dioxin in rats. Food Cosmet. Toxicol., 13:
    501-505.

    ALLEN, J.R., BARSOTTI, D.A., VAN MILLER, J.P., ABRAHAMSON, L.J., &
    LALICH, J.J. (1977) Morphological changes in monkeys consuming a diet
    containing low levels of 2,3,7,8-tetrachloro-dibenzo-p-dioxin. Food
    Cosmet. Toxicol., 15(5): 401-410.

    ALLEN, J.R., HARGRAVES, W.A., HSIA, M.T.S., & LIN, F.S.D. (1979a)
    Comparative toxicology of chlorinated compounds on mammalian species.
    Pharmacol. Ther., 7: 513-547.

    ALLEN, J.R., BARSOTTI, D.A., LAMBRECHT, L.K., & VAN MILLER, J.P.
    (1979b) Reproductive effects of halogenated aromatic hydrocarbons on
    nonhuman primates. Ann. N.Y. Acad. Sci., 320: 19-27.

    ANDERSSON, K., BOSSHARDT, H.P., BUSER, H.R., MARKLUND, S., & RAPPE, C.
    (1978) Chlorinated phenoxy acids and their dioxins: Chemistry-summary.
    Ecol. Bull., 27: 19-27.

    ARSTILA, A.U., REGGIANI, G., SORVARI, T.E., RAISANEN, S., & WIPF, H.K.
    (1981) Elimination of 2,3,7,8-tetrachlorodibenzo-p-dioxin in goat
    milk. Toxicol. Lett., 9: 215-219.

    ASHE, W.F. & SUSKIND, R.R. (1949) Clinical Report on four patients
    from Monsanto Chemical Company in Nitro, West Virginia, USA, 14 pp.
    (Submitted December 1949).

    ASHE, W.F. & SUSKIND, R.R. (1950) Clinical report on six patients
    from Monsanto Chemical Company in Nitro, West Virginia, USA, 16 pp.
    (Submitted April 1950).

    BAADER, E.W. & BAUER, H.J. (1951) Industrial intoxication due to
    pentachlorophenol. Ind. Med. Surg., 20(6): 286-290.

    BAARS, A.J., MUKHTAR, H., JANSEN, M., & BREIMER, D.D. (1982)
    Induction of rat hepatic glutathione-s-tranferase activities by
    2,3,7,8-tetrachlorodibenzo-p-dioxin, Oxford, New York, Pergamon
    Press, pp. 393-401 (Pergamon Series on Environmental Science, Vol. 5).

    BALK, J.L. & PIPER, W.N. (1984) Altered blood levels of
    corticosteroids in the rat after exposure to
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Biochem. Pharmacol., 33(15):
    2531-2534.

    BALL, L.M. & CHABRA, R.S. (1981) Intestinal absorption of nutrients in
    rats treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J.
    Toxicol. environ. Health, 8: 629-638.

    BANDIERA, S., FARRELL, K., MASON, G., KELLEY, M., ROMKES, M.,
    BANNISTER, R., & SAFE, S. (1984a) Comparative toxicities of the
    polychlorinated dibenzofuran (PCDF) and biphenyl (PCB) mixtures which
    persist in yusho victims. Chemosphere, 13(4): 507-512.

    BANDIERA, S., SAWYER, T., ROMKES, M., ZMUDZKA, B., SAFE, L., MASON,
    G., KEYS, B., & SAFE, S. (1984b) Polychlorinated dibenzofurans
    (PCDFs): Effects of structure on binding to the 2,3,7,8-TCDD cytosolic
    receptor protein, AHH induction and toxicity. Toxicology, 32:
    131-144.

    BANNISTER, R. & SAFE, S. (1987) Aroclor 1254 as a
    2,3,7,8-tetrachlorodibenzo-p-dioxin antagonist in C57BL/6J and DBA/2J
    mice - effects of enzyme induction. Toxicologist, 7: 160.

    BANNISTER, R., MASON, G., KELLEY, M., & SAFE, S. (1986) The effects of
    cytosolic receptor modulation on the AHH-inducing activity of
    2,3,7,8-TCDD. Chemosphere, 15: 1909-1911.

    BARSOTTI, D.A., ABRAHAMSON, L.J., & ALLEN, J.R. (1979) Hormonal
    alterations in female rhesus monkeys fed a diet containing
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Bull. environ. Contam.
    Toxicol., 21: 463-469.

    BARTELSON, F.D., HARRISON, D.D., & MORGAN, J.D. (1975) Field
    studies of wildlife exposed to TCDD contaminated soil, Florida,
    Airforce Armament Laboratory, Eglin Air Force Base (AFATL-TR-75-49).

    BASTOMSKY, C.H. (1977) Enhanced thyroxine metabolism and high uptake
    goiters in rats after a single dose of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Endocrinology, 101(1):
    292-296.

    BAUER, H., SCHULZ, K.H., & SPIEGELBERG, U. (1961) [Occupational
    poisoning in the manufacture of chlorophenol compounds.] Arch.
    Gewerbepathol. Gewerbehyg., 18: 538-555 (in German).

    BAUGHMAN, R.W. (1974) Tetrachlorodibenzo-p-dioxins in the
    environment. High resolution mass spectrometry of the picogram
    level, Cambridge, Massachussets, Harvard University (Thesis).

    BAUGHMAN, R.W. & MESELSON, M. (1973) An analytical method for
    detecting TCDD (dioxin): Levels of TCDD in samples from Vietnam.
    Environ. Health Perspect., 5: 27-35.

    BEALE, M.G., SHEARER, W.T., KARL, M.M., & ROBSON, A.M. (1977)
    Long-term effects of dioxin exposure. Lancet, 1: 748.

    BEATTY, P.W. & NEAL, R.A. (1977) Factors affecting the induction of
    DT-diaphorase by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem.
    Pharmacol., 27: 505-510.

    BEATTY, P.W., LEMBACH, K.J., HOLSCHER, M.A., & NEAL, R.A. (1975)
    Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on mammalian
    cells in tissue cultures. Toxicol. appl. Pharmacol., 31: 309-312.

    BEATTY, P.W., VAUGHN, W.K., & NEAL, R.A. (1978) Effect of alteration
    of rat hepatic mixed-function oxidase (MFO) activity on the toxicity
    of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicol. appl.
    Pharmacol., 45: 513-519.

    BECK, H., ECKART, K., KELLERT, M., MATHAR, W., RUHL, CH.-S., &
    WITTOWSKI, R. (1987) Levels of PCDF and PCDD in samples of human
    origin and food in the Federal Republic of Germany. Chemosphere,
    16: 1977-1982.

    BELL, R.A. & GARA, A. (1985) Synthesis and characterization of the
    isomers of polychlorinated dibenzofurans, tetra-through octa-. In:
    Keith, L.H., Rappe, C., & Choudhary, C.G., ed. Chlorinated dioxins
    and dibenzofurans in the total environment, Boston, Butterworth
    Publishers, pp. 3-16.

    BERGMAN, A., HAGMAN, A., JACOBSSON, S., JANSSON, B., & AHLMAN, M.
    (1984) Thermal degradation of polychlorinated alkanes. Chemosphere,
    13: 237-250.

    BERLIN, A., BURRATTA, A., & VAN DER VENNE, M.-T. (1967) Proceedings
    of the Expert Meeting on the Problems raised by TCDD-Pollution,
    Milan, Italy, Luxembourg, Commission of the European Communities, p.
    179.

    BERRY, D.L., ZACHARIAH, P.K., NAMKUNG, M.J., & JUCHAU, M.R. (1976)
    Transplacental induction of carcinogen-hydroxylating systems with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol., 36:
    569-548.

    BERRY, D.L., SLAGA, T.J., WILSON, N.M., ZACHARIAH, P.K., NAMKUNG,
    M.J., BRACKEN, W.M., & JUCHAU, M.R. (1977) Tranplacental induction of
    mixed-function oxygenases in extra-hepatic tissues by
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem. Pharmacol., 26:
    1383-1388.

    BERRY, D.L., SLAGA, T.J., DIGIOVANNI, J., & JUCHAU, M.R. (1979)
    Studies with chlorinated dibenzo-p-dioxins, polybrominated biphenyls,
    and poly chlorinated biphenyls in a two-stage system of mouse skin
    tumorigenesis: potent anticarcinogenic effects. Ann. N.Y. Acad. Sci.,
    320: 405-414.

    BERTONI, G., BROCCO, D., DI PALO, V., LIBERTI, A., POSSANZINI, M., &
    BRUNER, F. (1978) Gas chromatographic determination of 2,3,7,8-tetra
    chloro-dibenzodioxin in the experimental decon-tamination of Seveso
    soil by ultraviolet radiation. Anal. Chem., 50: 732-735.

    BIRNBAUM, L.S. (1986) Distribution and excretion of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin in congenic strains of mice which
    differ at the Ah locus. Drug Metab. Disp., 14(1): 34-40.

    BIRNBAUM, L.S., DECAD, G.M., & MATTHEWS, H.B. (1980) Disposition and
    excretion of 2,3,7,8-tetrachlorodibenzofuran in the rat. Toxicol.
    appl. Pharmacol., 55: 342-352.

    BIRNBAUM, L.S., DECAD, G.M., MATTHEWS, H.B., & MCCONNELL, E.E. (1981)
    Fate of 2,3,7,8-tetrachlorodibenzofuran in the monkey. Toxicol. appl.
    Pharmacol., 57: 189-196.

    BIRNBAUM, L.S., WEBER, H., HARRIS, M.W., LAMB, J.C., IV, & MCKINNEY,
    J.D. (1985) Toxic interaction of specific polychlorinated biphenyls
    and 2,3,7,8-tetrachlorodibenzo-p-dioxin: Increased incidence of cleft
    palate in mice. Toxicol. appl. Pharmacol., 77: 292-302.

    BIRNBAUM, L.S., HARRIS, M.W., MILLER, C.P., PRATT, R.M., & LAMB, J.C.
    (1986) Synergistic interaction of 2,3,7,8-tetra-chlorodibenzo-p-dioxin
    and hydro-cortisone in the induction of cleft palate in mice.
    Teratology, 33: 29-35.

    BIRNBAUM, L.S., HARRIS, M.W., BARNHART, E.R., & MORRISSEY, R.E. (1987)
    Teratogenicity of three polychlorinated dibenzo-furans in C57BL/6N
    mice. Toxicol. appl. Pharmacol., 90: 206-216.

    BLEIBERG, J., WALLEN, M., BRODKIN, R., & APPLEBAUM, I.L. (1964)
    Industrially acquired porphyria. Arch. Dermatol., 89: 793-797.

    BLOMHOFF, R., HELGERUD, P., RASMUSSEN, M., BERG, T., & NORUM, K.R.
    (1982) In vivo uptake of chylomicron (3H)retinyl ester by rat liver:
    Evidence for retinol transfer from parenchymal to nonparenchymal
    cells. Proc. Natl Acad. Sci. (USA), 79: 7326-7330.

    BOER, F.P., NEUMAN, M.A., VAN REMOORTERE, F.P., NORTH, P.P., & RINN,
    H.W., (1973) X-Ray diffraction studies of chlorinated
    dibenzo-p-dioxins. Chlorodioxins - origin and fate. Adv. Chem.
    Ser., 120, 14-25.

    BOMBICK, D.W., MATSUMURA, F., & MADHUKAR, B.V. (1984) TCDD
    (2,3,7,8-Tetrachlorodibenzo-p-dioxin) causes reduction in the low
    density lipoprotein (LDL) receptor activities in the hepatic plasma
    membrane of the guinea pig and rat. Biochem. biophys. Res.
    Commun., 118: 548-554.

    BOMBICK, D.W., MADHUKAR, B.V., BREWSTER, D.W., & MATSUMURA, F. (1985)
    TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin) causes increases in protein
    kinases particularly protein kinase C in the hepatic plasma membrane
    of the rat and the guinea pig. Biochem. biophys. Res. Commun.,
    127(1): 296-302.

    BONACCORSI, A., DI DOMENICO, A., FANELLI, R., MERLI, F., MOTTA, R.,
    VANZATI, R., & ZAPPONI, G.A. (1984) The influence of soil particle
    adsorption on 2,3,7,8-tetrachlorodibenzo-p-dioxin biological uptake in
    the rabbit. Arch. Toxicol., 7: 431-434.

    BOTRE, C., MEMOLI, A., & ALHAIQUE, F. (1978) TCDD solubilization and
    photodecomposition in aqueous solutions. Environ. Sci. Technol.,
    12: 335-336.

    BOWES, G.W., MULVIHILL, M.J., DECAMP, M.R., & KENDE, A.S. (1975) Gas
    chromatographic characteristics of authentic chlorinated
    dibenzofurans; identification of two isomers in American and Japanese
    polychlorinated biphenyls. J. agric. food Chem., 23(6): 1222-1223.

    BRADLAW, J.A. & CASTERLINE, J.L., Jr (1979) Induction of enzyme
    activity in cell culture: a rapid screen for detection of planar
    polychlorinated organic compounds. J. Assoc. Off. Anal. Chem.,
    62(4): 904-916.

    BRADLAW, J.A., GARTHOFF, L.H., HURLEY, N.E., & FIRESTONE, D. (1980)
    Comparative induction of aryl hydrocarbon hydroxylase activity in
    vitro by analogues of dibenzo-p-dioxin. Food Cosmet. Toxicol., 18:
    627-635.

    BRAUN, W. (1955) [Chloracne. Monographs with the Journal
    Berufsdermatosen,] Aulendorf i. Wurtt., Editio Cantor, Vol. 1. (in
    German).

    BREWSTER, D.W., MADHUKAR, B.V., & MATSUMURA, F. (1982) Influence of
    2,3,7,8-TCDD on the protein composition of the plasma membrane of
    hepatic cells from the rat. Biochem. biophys. Res. Commun., 107:
    68-74.

    BRONZETTI, G., BAUER, C., CORSI, C., DEL CARRATORE, R., NIERI, R., &
    PAOLINI, M. (1983) Mutagenicity study of TCDD and ashes from urban
    incinerator "in vitro" and "in vivo" using yeast D7 strain.
    Chemosphere, 12(4/5): 549-553.

    BROWN, E.F. & MORGAN, A.F. (1948) The effect of vitamin A deficiency
    upon the nitrogen metabolism of the rat. J. Nutr., 35: 425-438.

    BUMB, R.R., CRUMMETT, W.B., CUTIE, S.S., GLADHILL, R.H., VAGEL, R.O.,
    LAMPAVSKI, L.L., LUOMA, E.V., MILLER, D.L., NESTRIDE, T.J., SHADOFF,
    L.A., STEHL, R.H., & WOODS, J.S., (1980) Trace chemistries of fire: A
    source of chlorinated dioxins. Science, 210: 385-390.

    BURKHARD, L.P. & KUEHL, D.W. (1986) N-Octanol/water partition
    coefficients by reverse phase liquid chromatography/mass spectrometry
    for eight tetrachlorinated planar molecules. Chemosphere, 15:
    163-167.

    BUS, J.S. & GIBSON, J.E. (1979) Lipid peroxidation and its role in
    toxicology. Rev. Biochem. Toxicol., 1: 125-149.

    BUSER, H.R. (1975) Polychlorinated dibenzo-p-dioxins. Separation and
    identification of isomers by gas chromotagrophy mass spectrometry. J.
    Chromatogr., 114: 95-108.

    BUSER, H.R. (1979) Formation of polychlorinated dibenzofurans (PCDFs)
    and dibenzo-p-dioxins (PCDDs) from the pyrolysis of chlorobenzenes.
    Chemosphere, 6: 415-424.

    BUSER, H.R. & BOSSHARDT, H.P. (1976) Determination of polychlorinated
    dibenzo-p-dioxins and dibenzofurans in commercial pentachlorophenols
    by combined gas chromatography - mass spectrometry. J. Assoc. Off.
    Anal. Chem., 59: 562-569.

    BUSER, H.R. & BOSSHARDT, H.P. (1978) [Polychlorinated
    dibenzo-p-dioxin, dibenzofuran and benzene in the ashes from municipal
    and industrial incinerators.] Mitt. Geb. Lebensm. Hyg., 69:
    191-199 (in German).

    BUSER, H.R. & RAPPE, C. (1978) Identification of substitution patterns
    in polychlorinated dibenzo-p-dioxins (PCDDs) by mass spectrometry.
    Chemosphere, 7: 199-211.

    BUSER, H.R. & RAPPE, C. (1979) Formation of polychlorinated
    dibenzofurans (PCDFs) from the pyrolysis of individual PCB isomers.
    Chemosphere, 3: 157-174.

    BUSER, H.R. & RAPPE, C. (1980) High-resolution gas chromato-graphy of
    the 22 tetrachlorodibenzo-p-dioxin isomers. Anal. Chem., 52:
    2257-2262.

    BUSER, H.R. & RAPPE, C. (1984) Isomer-specific separation of
    2,3,7,8-substituted polychlorinated dibenzo-p-dioxins by
    high-resolution gas chromatography/mass spectrometry. Anal. Chem.,
    56: 442-448.

    BUSER, H.R., BOSSHARDT, H.P., & RAPPE, C. (1978a) Formation of
    polychlorinated dibenzofurans (PCDFs) from the pyrolysis of PCBs.
    Chemosphere, 1: 109-119.

    BUSER, H.R., BOSSHARDT, H.P., & RAPPE, C. (1978b) Identification of
    polychlorinated dibenzo-p-dioxin isomers found in fly ash.
    Chemosphere, 7(2): 165-172.

    BUSER, H.R., BOSSHARDT, H.P., RAPPE, C., & LINDAHL, R. (1978c)
    Identification of polychlorinated dibenzofuran isomers in fly ash and
    PCB pyrolysis. Chemosphere, 5: 419-429.

    BUSER, H.R., RAPPE, C., & GARA, A. (1978d) Polychlorinated
    dibenzofurans (PCDFs) found in Yosho oil and in used Japanese PCB.
    Chemosphere, 5: 439-449.

    BUSER, H.R., RAPPE, C., & BERGQVIST, P.-A. (1985) Analysis of
    polychlorinated dibenzofurans, dioxins, and related compounds in
    environmental samples. Environ. Health Perspect., 60: 293-302.

    BUU-HOI, N.P., HIEN, D.P., SAINT-RUF, G., & SERVOIN-SIDOINE, J.
    (1971a) Proprits cancromimtiques de la ttrachloro-2,3,7,8
    dibenzo-p-dioxin. C. R. Acad. Sci. Paris, 272: 1447-1450.

    BUU-HOI, N.P., SAINT-RUF, G., BIGOT, P., & MANGANE, M. (1971b)
    Prparation, proprits et identification de la "dioxine"
    (ttrachloro-2,3,7,8-dibenzo-p-dioxine) dans les pyrolysats de
    dfoliants  base d'acide trichloro-2,4,5 phnoxyactique et de ses
    esters et des vgtaux contamins. C. R. Acad. Sci. Paris, 273:
    708-711.

    BUU-HOI, N.P., CHANH, P.H., SESQUE, G., AZUM-GELADE, M.C., &
    SAINT-RUF, G. (1972a) Organs as targets of "dioxin"
    (2,3,7,8-tetrachlorodibenzo-p-dioxin) intoxication.
    Natur-wissenschaften, 59: 174-175.

    BUU-HOI, N.P., CHANH, P.H., SESQUE, G., AZUM-GELADE, M.C., &
    SAINT-RUF, G. (1972b) Enzymatic functions as targets of the toxicity
    of "dioxin" (2,3,7,8-tetrachlorodibnzo-p-dioxin.
    Naturwissenschaften, 59: 173-174.

    CANTONI, L., SALMONA, M., & RIZZARDINI, M. (1981) Porphyrogenic effect
    of chronic treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin in
    female rats. Dose-effect relationship following urinary excretion of
    porphyrins. Toxicol. appl. Pharmacol., 57: 156-163.

    CANTONI, L., DAL FIUME, D., FERRAROLI, A., SALMONA, M., & RUGGIERI, R.
    (1984a) Different susceptibility of mouse tissues to porphyrogenic
    effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. Lett., 20:
    201-210.

    CANTONI, L., DAL FIUME, D., RIZZARDINI, M., & RUGGIERI, R. (1984b) In
    vitro inhibitory effect on porphyrinogen carboxylase of liver extracts
    from TCDD treated mice. Toxicol. Lett., 20: 211-217.

    CARLSTEDT-DUKE, J.M.B. (1979) Tissue distribution of the receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Cancer Res., 39:
    3172-3176.

    CARLSTEDT-DUKE, J.M.B, ELFSTROEM, G., HOEGBERG, B., & GUSTAFSSON,
    J.-A. (1979) Oncogeny of the rat hepatic receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin and its endocrine independence.
    Cancer Res., 39: 4653-4656.

    CARLSTEDT-DUKE, J.M.B., HARNEMO, U.-B., HOEGBERG, B., & GUSTAFSSON,
    J.-A. (1981) Interaction of the hepatic receptor protein for
    2,3,7,8-tetrachlorodibenzo-p-dioxin with DNA. Biochim. Biophys.
    Acta, 672: 131-141.

    CARLSTEDT-DUKE, J.M.B., KURL, R., POELLINGER, L., GILLNER, M.,
    HANSSON, L.-A., TOFTGARD, R., HOEGBERG, B., & GUSTAFSSON, J.-A. (1982)
    The detection and function of the cytosolic receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related cocarcinogens.
    In: Hutzinger, O., ed. Chlorinated dioxins and related compounds.
    Impact on the environment, Oxford, New York, Pergamon Press, pp.
    355-365

    CARTER, C.D., KIMBROUGH, R.D., LIDDLE, J.A., CLINE, R.E., ZACH, M.M.,
    Jr, BARTHEL, W.F., KOEHLER, R.E., & PHILLIPS, P.E. (1975)
    Tetrachlorodibenzo-dioxin: An accidental poisoning episode in horse
    arenas. Science, 188: 738-740.

    CAVALLARO, A., BANDI, G., INVERNIZZI, G., LUCIANI, L., MONGINI, E., &
    GORNI, G. (1982) Negative ion chemical ionization ms as a structure
    tool in the determination of small amounts of PCDD and PCDF. In:
    Hutzinger, O., ed. Chlorinated dioxins and related compounds,
    Oxford, New York, Pergamon Press, Vol. 5, pp. 55-65.

    CECIL, H.C., HARRIS, S.J., BITMAN, J., & FRIES, G.F. (1973)
    Polychlorinated biphenyl-induced decrease in liver vitamin A in
    Japanese Quail and rats. Bull. environ. Contam. Toxicol., 9:
    179-185.

    CHANG, K.J., HSIEH, K.H., LEE, T.P., TANG, S.Y., & TUNG, T.C. (1981)
    Immunologic evaluation of patients with polychlorinated biphenyl
    poisoning: determination of lymphocyte subpopulations. Toxicol.
    appl. Pharmacol., 61: 58-63.

    CHANG, K.J., HSIEH, K.H., LEE, T.P., & TUNG, T.C. (1982a) Immunologic
    evaluation of patients with polychlorinated biphenyl poisoning:
    determination of phagocyte Fc and complement receptors. Environ. Res.,
    28: 329-334.

    CHANG, K.J., HSIEH, K.H., TANG, S.Y., TUNG, T.C., & LEE, T.P. (1982b)
    Immunologic evaluation of patients with polychlorinated biphenyl
    poisoning: evaluation of delayed-type skin hypersensitive response and
    its relation to clinical studies. J. Toxicol. environ. Health,
    9: 217-223.

    CHAPMAN, D.E. & SCHILLER, C.M. (1985) Dose-related effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in C57BL/6J and DBA/2J
    mice. Toxicol. appl. Pharmacol., 78: 147-157.

    CHASTAIN, J.E. & PAZDERNIK, T.L. (1985)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin (TCDD)-induced immunotoxicity.
    Int. J. Immunopharmacol., 7(6): 849-856.

    CHEN, P.H. & HITES, R.A. (1983) Polychlorinated biphenyls and
    dibenzofurans retained in the tissues of a deceased patient with
    Yu-Cheng in Taiwan. Chemosphere, 12: 1507-1516.

    CHEN, P.H., GAW, J.M., WONG, C.K., & CHEN, C.J. (1980) Levels and gas
    chromatographic patterns of polychlorinated biphenyls in the blood of
    patients after PCB poisoning in Taiwan. Bull. environ. Contam.
    Toxicol., 25: 325-329.

    CHEN, P.H., CHANG, K.T., & LU, Y.D. (1981) Polychlorinated biphenyls
    and polychlorinated dibenzofurans in the toxic rice-bran oil that
    caused PCB poisoning in Taichung. Bull. environ. Contam. Toxicol.,
    26, 489-495.

    CHEN, P.H., WONG, C.K., RAPPE, C., & NYGREN, M. (1985) Polychlorinated
    biphenyls, dibenzofurans and quaterphenyls in toxic rice-bran oil and
    in the blood and tissues of patients with PCB poisoning (Yu-Cheng) in
    Taiwan. Environ. Health Perspect., 59: 59-65.

    CHEUNG, M.O., GILBERT, E.F., & PETERSON, R.E. (1981) Cardiovascular
    teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the chick
    embryo. Toxicol. appl. Pharmacol., 61: 197-204.

    CHRISTIAN, B.J., INHORN, S.L., & PETERSON, R.E. (1986a) Relationship
    of the wasting syndrome to lethality in rats treated with
    2,3,7,8-tetrachloro-dibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    82: 239-255.

    CHRISTIAN, B.J., MENAHAN, L.A., & PETERSON, R.E. (1986b) Intermediary
    metabolism of the mature rat following
    2,3,7,8-tetrachlorodibenzo-p-dioxin treatment. Toxicol. appl.
    Pharmacol., 83: 360-378.

    CLARK, D.A., GAULDIE, J., SZEWCZUK, M.R., & SWEENEY, G. (1981)
    Enhanced suppressor cell activity as a mechanism of immunosuppression
    by 2,3,7,8-tetrachlorodibenzo-p-dioxin (41275). Proc. Soc. Exp. Biol.
    Med., 168: 290-299.

    CLARK, D.A., SWEENEY, G., SAFE, S., HANCOCK, E., KILBURN, D.G., &
    GAULDIE, J. (1983) Cellular and genetic basis for suppression of
    cytotoxic T-cell generation by haloaromatic hydrocarbons.
    Immunopharmacology, 6: 143-153.

    CLEMENT, R.E., TOSINE, H.M., & ALI, B. (1985) Levels of
    polychlorinated dibenzo-p-dioxins and dibenzofurans in wood burning
    stoves and fireplaces. Chemosphere, 14: 815-819.

    COCCIA, P., CROCI, T., & MANARA, L. (1981) Less TCDD persists in liver
    2 weeks after a single dose to mice fed chow with added charcoal or
    cholic acid. Br. J. Pharmacol., 72: 181-182.

    COCHRANE, W.P., SINGH, J., MILES, W., WAKEFORD, B., & SCOTT, J. (1981)
    Analysis of technical and formulated products of 2,4-dichlorophenoxy
    acid for the presence of chlorinated dibenzo-p-dioxins. In: Hutzinger,
    O., Frei, R.W., Merian, E., & Pocchiari, F., ed. Impact of
    chlorinated dioxins and related compounds on the environment,
    Oxford, New York, Pergamon Press, pp. 209-213.

    COHEN, G.M., BRACKEN, W.M., IYER, R.P., BERRY, D.L., SELKIRK, J.K., &
    SLAGA, T.J. (1979) Anticarcinogenic effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on benzo(a)pyrene and
    7,12-dimethylbenz(a)anthracene tumor initiation and its relationship
    to DNA binding. Cancer Res., 39: 4027-4033.

    CONAWAY, C.C. & MATSUMURA, F. (1975) Alteration of cellular
    utilization of thymidine by TCDD (2,3,7,8-tetrachlorodibenzo-p
    -dioxin). Bull. environ. Contam. Toxicol., 12: 52-56.

    CONSTABLE, J.D. & HATCH, M.C. (1985) Reproductive effects of herbicide
    exposure in Vietnam: Recent studies by the Vietnamese and others.
    Teratog. Carcinog. Mutagen., 5: 231-250. 
    COOK, R.R. (1981) Dioxin, chloracne, and soft tissue sarcoma.
    Lancet, 1: 618-619.

    COOK, R.R., TOWNSEND, J.C., OTT, M.G., & SILVERSTEIN, L.G. (1980)
    Mortality experience of employees exposed to
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J. occup. Med., 22:
    530-532.

    COURTNEY, K.D. (1976) Mouse teratology studies with
    chlorodibenzo-p-dioxins. Bull. environ. Contam. Toxicol., 16:
    674-681.

    COURTNEY, K.D. & MOORE, J.A. (1971) Teratology studies with
    2,4,5-trichlorophenoxy-acetic acid and
    2,3,7,8-tetrachloro-dibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    20: 396-403.

    COURTNEY, K.D., GAYLOR, D.W., HOGAN, M.D., FALK, H.L., BATES, R.R., &
    MITCHEL, I. (1970) Teratogenic evaluation of 2,4,5-T. Science,
    168: 864-866.

    COURTNEY, K.D., PUTNAM, J.P., & ANDREWS, J.E. (1978) Metabolic studies
    with TCDD (Dioxin) treated rats. Arch. environ. Contam. Toxicol.,
    7: 385-396.

    COWARD, K.H. (1947) The determination of vitamin A. In: The
    biological standardisation of the vitamins, London, Baillire,
    Tindall and Cox, pp. 23-58.

    CROSBY, D.G. & WONG, A.S. (1976) Photochemical generation of
    chlorinated dioxins. Chemosphere, 5: 327-332.

    CROSBY, D.G. & WONG, A.S. (1977) Environmental degradation of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Science, 195:
    1337-1338.

    CROSBY, D.G., WONG, A.S., PLIMMER, J.R., & WOOLSON, E.A. (1971)
    Photodecomposition of chlorinated dibenzo-p-dioxins. Science, 173:
    748-749.

    CROSBY, D.G., MOILANEN, K.W., & WONG, A.S. (1973) Environmental
    generation and degradation of dibenzodioxins and dibenzofurans.
    Environ. Health Perspect., 5: 259-266.

    CROW, K.D. (1970) Chloracne. A critical review including a comparison
    of two series of cases of acne from chlornaphtalene and pitch fumes.
    Trans. St. John's Hosp. Dermatol. Soc., 56: 79-99.

    CRUMMETT, W.B. & STEHL, R.H. (1973) Determination of chlorinated
    dibenzo-p-dioxins and dibenzofurans in various materials. Environ.
    Health Perspect., 5: 15-25.


    CRUMMETT, W.B., NESTRICK, T.J., & LAMPARSKI, L.L., (1985) Analytical
    methodology for the determination of PCDDs in environmental samples:
    an overview and critique. In: Kamrin, M.A. & Rodgers, P.W., ed.
    Dioxins in the environment, Washington, DC, Hemisphere Publishing,
    pp. 57-83.

    CUNNINGHAM, H.M. & WILLIAMS, D.T. (1972) Effect of
    tetrachlorodibenzo-p-dioxin on growth rate and the synthesis of lipids
    and proteins in rats. Bull. environ. Contam. Toxicol., 7: 45-51.

    CZUCZWA, J.M. & HITES, R.A. (1985) Historical record of
    polychlorinated dioxins and furans in Lake Huron sediments. In: Keith,
    L.H., Rappe, C., & Choudhary, G., ed. Chlorinated dioxins and
    dibenzofurans in the total environment, Stoneham, Maine, Butterworth
    Publishers, pp. 59-63.

    DALDERUP, L.M. (1974a) Safety measures for taking down buildings
    contaminated with toxic material I. T. Soc. Geneeskd., 52: 582-588.

    DALDERUP, L.M. ((1974b) Safety measures for taking down buildings
    contaminated with toxic material II. T. Soc. Geneeskd., 52:
    616-623.

    D'ARGY, R., HASSOUN, E., & DENCKER, L. (1984) Teratogenicity of TCDD
    and the congener 3,3',4,4'-tetrachloroazoxybenzene in sensitive and
    nonsensitive mouse strains after reciprocal blastocyst transfer.
    Toxicol. Lett., 21: 197-202.

    DAVISON, K.L. & COX, J.H. (1976) Methoxychlor effects on hepatic
    storage of vitamin A in rats. Bull. environ. Contam. Toxicol., 16:
    145-148.

    DECAD, G.M., BIRNBAUM, L.S., & MATTHEWS, H.B. (1981a)
    2,3,7,8-tetrachlorodibenzofuran tissue distribution and excretion in
    guinea pigs. Toxicol. appl. Pharmacol., 57: 231-240.

    DECAD, G.M., BIRNBAUM, L.S., & MATTHEWS, H.B. (1981b) Distribution and
    excretion of 2,3,7,8-tetrachlorodibenzofuran in C57BL/6J and DBA/2J
    mice. Toxicol. appl. Pharmacol., 59: 564-573.

    DECAPRIO, A.P., MCMARTIN, D.N., SILKWORTH, J.B., REJ, R., PAUSE, R.,
    & KAMINSKY, L.S. (1983) Subchronic oral toxicity in guinea pigs of
    soot from a polychlorinated biphenyl-containing transformer fire.
    Toxicol. appl. Pharmacol., 68: 308-322.


    DECAPRIO, A.P., MCMARTIN, D.N., O'KEEFE, P.W., REJ, R., SILKWORTH,
    J.B., & KAMINSKY, L.S. (1986) Subchronic oral toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in the guinea pig: comparisons
    with a PCB-containing transformer fluid pyrolysate. Fundam. appl.
    Toxicol., 6: 454-463.

    DEITRICH, R.A., BLUDEAU, P., ROPER, M., & SCHMUCK, J. (1978) Induction
    of aldehyde dehydrogenases. Biochem. Pharmacol., 27: 2343-2347.

    DENCKER, L. & PRATT, R.M. (1981) Association between the presence of
    the Ah receptor in embryonic murine tissues and sensitivity to
    TCDD-induced cleft palate. Teratog. Carcinog. Mutagen., 1:
    399-406.

    DENCKER, L., HASSOUN, E., D'ARGY, R., & ALM, G. (1985) Fetal thymus
    organ culture as an in vitro model for the toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin and its congeners. Mol.
    Pharmacol., 27: 133-140.

    DENISON, M.S., VELLA, L.M., & OKEY, A.B. (1986a) Structure and
    function of the Ah receptor for 2,3,7,8-tetrachlorodibenzo-p-dioxin.
    Species difference in molecular properties of the receptors from mouse
    and rat hepatic cytosols. J. biol. Chem., 261: 3987-3995.

    DENISON, M.S., WILKINSON, C.F., & OKEY, A.B. (1986b) Ah receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin: comparative studies in mammalian
    and nonmammalian species. Chemosphere, 15: 1665-1672.

    DENISON, M.S., HARPER, P.A., & OKEY, A.B. (1986c) Ah receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Codistribution of unoccupied
    receptor with cytosolic marker enzymes during fractionation of mouse
    liver, rat liver and cultured Hepa-1c1 cells. Eur. J. Biochem.,
    155: 223-229.

    DENMARK (1984) [Formation and emission of dioxins especially in
    connection with waste incineration: supplement,] Copenhagen,
    Miljstyrelsen (in Danish).

    DICKINS, M., SEEFELD, M.D., & PETERSON, R.E. (1981) Enhanced liver DNA
    synthesis in partially hepatectomized rats pre-treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    58: 389-398.

    DIDOMENICO, A., SILANO, V., VIVIANO, G., & ZAPPONI, G. (1980)
    Accidental release of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) at
    Seveso, Italy. II. TCDD distribution in the soil surface layer.
    Ecotoxicol. environ. Saf., 4: 298-320.


    DIGIOVANNI, J., VIAJE, A., BERRY, D.L., SLAGA, T.J., & JUCHAU, M.R.
    (1977) Tumor-initiating ability of
    2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD) and arochlor 1254 in the
    two-stage system of mouse skin carcinogenesis. Bull. environ.
    Contam. Toxicol., 18: 552-557.

    DIGIOVANNI, J., BERRY, D.L., JUCHAU, M.R., & SLAGA, T.J. (1979)
    2,3,7,8-Tetra-chlorodibenzo-p-dioxin: potent anti-carcinogenic
    activity in CD-1 mice. Biochem. biophys. Res. Commun., 86:
    577-584.

    DIGIOVANNI, J., BERRY, D.L., GLEASON, G.L., KISHORE, G.S., & SLAGA,
    T.J. (1980) Time dependent inhibition by
    2,3,7,8-tetra-chlorodibenzo-p-dioxin of skin tumorogenesis with
    polycyclic hydrocarbons. Cancer Res., 40: 1580-1587.

    DOBLES, A.J. & GRANT, C. (1979) Photolysis of highly chlorinated
    dibenzo-p-dioxins by sunlight. Nature (Lond.), 278: 162-165.

    DOHMEIER, H.J. & JANSON, E. (1983) [Killing flies and people:
    Dioxin - the poison of Seveso and Viet Nam and how we come into
    daily contact with it.] Reinbek, Aktuell Rororo, p. 25.

    DONOVAN, J.W., MACLENNAN, R., & ADENA, N. (1984) Vietnam service and
    the risk of congenital anomalies: a case-control study. Med. J.
    Aust., 140: 394-397.

    DOYLE, B.W., DRUM, D.A., & LAUDER, J.D. (1985) The smoldering question
    of hospital wastes. Pollut. Eng., 17: 35-39.

    DOYLE, E.A. & FRIES, G.F. (1986) Induction of aryl hydrocarbon
    hydroxylase by chlorinated dibenzofurans in rats. Chemosphere, 15:
    1745-1748.

    DRINKER, C.K., FIELD WARREN, M., & GRANVILLE, A. (1937) The problem of
    possible systemic effects from certain chlorinated hydrocarbons. J.
    ind. Hyg. Toxicol., 19: 283-311.

    DUGOIS, P. & COLOMB, L. (1956) Acn chlorique au 2-4-5
    trichlorophnol. J. Md. Lyon, 88: 446-447.

    DUGOIS, P. & COLOMB, L. (1957) Remarques sur l'acn chlorique. A
    propos d'une closion de cas provoqus par la prparation du
    2,4,5-trichlorophnol. J. Md. Lyon, 38: 899-903.

    DUGOIS, P., MARECHAL, J., & COLOMB, L. (1958) Acn chlorique au
    2,4,5-tri-chlorophnol. Arch. Mal. prof. Hyg. Toxicol. ind., 19:
    626-627.


    DUGOIS, P., AMBLARD, P., AIMARD, M., & DESHORS, G. (1967) Acn
    chlorique collective et accidentelle d'un type nouveau. Bull. Soc.
    Dermatol., 75: 260-261.

    DUVERNE, J., THIVOLET, J., & BERARD, J. (1964) Acn chlorique profuse
    avec panchement pleural rcidivant chez un sujet  ractions
    tuberculiniques ngatives. Bull. Soc. Fr. Dermatol. Syphiligr.,
    71: 649-652.

    EATON, D.L. & KLAASSEN, C.D. (1979) Effects of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin, kepone, and polybrominated
    biphenyls on transport systems in isolated rat hepatocytes. Toxicol.
    appl. Pharmacol., 51: 137-144.

    EDULJEE, G.H., ATKINS, D.H.F., & EGGLETON, A.E. (1986) Observations
    and assessment relating to incineration of chlorinated chemical
    wastes. Chemosphere, 15: 1577-1584.

    ELDER, D.G., LEE, G.B., & TOVEY, J.A. (1978) Decreased activity of
    hepatic uroporphyrinogen decarboxylase in sporadic porphyria cutanea
    tarda. New Engl. J. Med., 299: 274-278.

    ELDER, G.H. & SHEPPARD, D.M. (1982) Immunoreactive uroporphyrinogen
    decarboxylase is unchanged in porphyria caused by TCDD and
    hexachlorobenzene. Biochem. biophys. Res. Commun., 109: 113-120.

    ELDER, G.H., EVANS, J.O., & MATLIN, S.A. (1976) The effect of
    porhyrogenic compound, hexachlorobenzene, on the activity of hepatic
    uroporphyrinogen decarboxylase in the rat. Clin. Sci. mol. Med.,
    51: 71-80.

    ERICKSON, J.D., MULINARE, J., MCCLAIN, P.W., FITCH, T.G., JAMES, L.M.,
    MCCLEARN, A.B., & ADAMS, M.J. (1984) Vietnam veterans' risks for
    fathering babies with birth defects. J. Am. Med. Assoc., 252:
    903-937.

    ESPOSITO, G., FAELLI, A., & CAPRARO, V. (1967) Metabolism and
    transport phenomena in isolated intestine of normal and semi-starved
    rats. Arch. int. Physiol. Biochim., 75: 601-608.

    ESPOSITO, M.P., TIERNAN, T.O., & DRYDEN, F.E. (1980) Dioxins,
    Cincinnati, Ohio, US Environmental Protection Agency, Office of
    Research and Development (EPA-600/2-80-197).

    FACCHETTI, S., BALASSO, A., FICHTNER, C., FRARE, G., LEONI, A., MAURI,
    C., & VASCONI, M. (1986) Studies on the absorption of TCDD by plant
    species. In: Rappe, C., Choudhary, G., & Keith, L.H., ed. Chlorinated
    dioxins and dibenzofurans in perspective, Chelsea, Michigan, Lewis
    Publishers, pp. 225-239.


    FAHRIG, R., NILSSON, C.-A., & RAPPE, C. (1978) Genetic activity of
    chlorophenols and chlorophenol impurities. Environ. Sci. Res., 12:
    325-338.

    FAITH, R.E. & LUSTER, M.I. (1979) Investigations on the effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on parameters of various
    immune functions. Ann. N.Y. Acad. Sci., 320: 564-571.

    FAITH, R.E. & MOORE, J.A. (1977) Impairment of thymus-dependent immune
    functions by exposure of the developing immune system to
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J. Toxicol. environ.
    Health, 3: 451-464.

    FARA, G.M., DEL CORNO, G., BONETTI, F., CARAMASHI, F., DARDANONI, L.,
    FAVARETTI, C., GIAMBELLUCA, S.E., MARNI, E., MOCARELLI, P.,
    MONTESARCHIO, E., PUCCINELLI, V., & VOLPATO, C. (1976) Chloracne after
    release of TCDD at Seveso, Italy. In: Hutzinger, O., Frei, R., Merian,
    E., & Pocchiari, F. ed. Chlorinated dioxins and related compounds.
    Impact on the environment, Oxford, New York, Pergamon Press, pp.
    545-559.

    FARRELL, K. & SAFE, S. (1986) 2,3,7,8-tetrachlorodibenzo-p-dioxin:
    Relationship between toxicity and the induction of aryl hydrocarbon
    hydroxylase and ornithine decarboxylase. Chemosphere, 15:
    1971-1976.

    FEDERAL REGISTER (1980) Storage and disposal of waste material:
    Prohibition of disposal of tetrachlorodibenzo-p-dioxin. Fed. Reg.,
    45(98): 32676.

    FILIPPINI, G., BORDO, B., CRENNA, P., MASSETTO, N., MUSICCO, M., &
    BOERI, R. (1981) Relationship between clinical and
    electrophysiological findings and indicators of heavy exposure to
    2,3,7,8-tetrachlorodibenzo-dioxin. Scand. J. Work Environ. Health,
    7: 257-262.

    FINGERHUT, M.A., HALPERIN, W.E., HONCHAR, P.A., SMITH, A.B., GROTH,
    D.H., & RUSSELL, W.O. (1984) Review of exposure and pathology data for
    seven cases reported as soft tissue sarcoma among persons
    occupationally exposed to dioxin-contaminated herbicides. In:
    Lowrance, W.W., ed. Public health risks of the dioxins. Proceedings
    of a Symposium held at the Rockefeller University, New York, 19-20
    October, 1983, Los Altos, California, William Kaufmann, pp. 187-203.

    FIRESTONE, D. (1973) Etiology of chick edema disease. Environ.
    Health Perspect., 5: 59-66.


    FIRESTONE, D., (1978) The 2,3,7,8-tetrachlorodibenzo-para-dioxin
    problem: A review. Proceedings of a Conference on Chlorinated
    Phenoxyacids and their Dioxins, Stockholm, 1977. Ecol. Bull., 27:
    39-52.

    FIRESTONE, D., CLOWER, M., Jr, BORSETTI, A.P., TESKE, R.H., & LONG,
    P.E. (1979) Polychlorodibenzo-p-dioxin and pentachloro-phenol residues
    in milk and blood of cows fed technical pentachlorophenol. J. agric.
    food Chem., 27: 1171-1177.

    FIRESTONE, D., NIEMANN, R.A., SCHNIDER, L.F., GRIDLEY, J.R., & BROWN,
    D.E., (1986) Dioxin residues in fish and other food. In: Rappe, C.,
    Choudhary, G., & Keith, L., ed. Chlorinated dioxins and
    dibenzofurans in perspective, Chelsea, Michigan, Lewis Publishers.

    FLICK, D.F., FIRESTONE, D., & HIGGINBOTHAM, G.R. (1972) Studies of the
    chick edema disease. 9. Response of chicks fed or singly administered
    synthetic edema-producing compounds. Poult. Sci., 51: 2026-2034.

    FLICK, D.F., FIRESTONE, D., RESS, J., & ALLEN, J.R. (1973) Studies of
    the chick edema disease. 10. Toxicity of chick edema factors in the
    chick, chick embryo, and monkey. Poult. Sci., 52: 1637-1641.

    FORTH, W. (1977) [The Seveso Disaster.] Dtsch. Arztebl., 44:
    2617-2626 (in German).

    FOWLER, B.A., LUCIER, G.W., BROWN, H.W., & MCDANIEL, O.S. (1973)
    Ultrastructural changes in rat liver cells following a single oral
    dose of TCDD. Environ. Health Perspect., 5: 141-148.

    FRAENKEL, C. (1902) [Reports of associations and congresses. Medical
    Association of Hall: Meeting of 23 October.] Mnchener Med.
    Wochenschr., October: 39-41.

    FREEDMAN, H.J., PARKER, N.B., MARINELLO, A.J., GURTOO, H.L., &
    MINOWADA, J. (1979) Induction, inhibition and biological properties of
    aryl hydrocarbon hydroxylase in a stable human B-lymphocyte cell line,
    RPMI-1788. Cancer Res., 39: 4612-4619.

    FREEMAN, R.A., SCHROY, J.M., HILEMAN, F.D., & NOBLE, R.W. (1986)
    Environmental mobility of 2,3,7,8-TCDD and companion chemicals in a
    roadway soil matrix. In: Rappe, C., Choudhary, G., & Keith, L., ed.
    Chlorinated dioxins and dibenzofurans in perspective, Chelsea,
    Michigan, Lewis Publishers, pp. 171-183.


    FRIES, G.F. & MARROW, G.S. (1975) Retention and excretion of
    2,3,7,8-tetrachlorodibenzo-p-dioxin by rats. J. agric. food Chem.,
    23: 265-269.

    FRIESEN, K.J., SARNA, L.P., & WEBSTER, G.R.B. (1985) Aqueous
    solubility of polychlorinated dibenzo-p-dioxins determined by high
    pressure liquid chromatography. Chemosphere, 14: 1267-1274.

    FUCHS, E. & GREEN, H. (1981) Regulation of terminal differentiation of
    cultured human keratinocytes by vitamin A. Cell, 25: 617-625.

    FUNG, D., BOYD, R.K., SAFE, S., & CHITTIM, B.G. (1985) Gas
    chromatographic/mass spectrometric analysis of specific isomers of
    polychlorodibenzofurans. Biomed. mass Spectrom., 12: 247-253.

    FURST, P., MEEMKEN, H.-A. KRUGER, Chr., & GROEBEL, W. (1987)
    Polychlorinated dibenzodioxins and dibenzofurans in human milk samples
    from Western Germany. Chemosphere, 16: 1983-1988.

    FURUHASHI, N., KURL, R.N., WONG, J., & VILLEE, C.A. (1986) A cytosolic
    binding protein for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the
    uterus and deciduoma of rats. Pharmacology, 33: 110-120.

    GALLO, M.A., HESSE, E.J., MACDONALD, G.J., & UMBREIT, T.H. (1986)
    Interactive effects of estradiol and
    2,3,7-8-tetra-chlorodibenzo-p-dioxin on hepatic cytochrome P-450 and
    mouse uterus. Toxicol. Lett., 32: 123-132.

    GASIEWICZ, T.A. & NEAL, R.A. (1979)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin tissue distribution, excretion,
    and effects on clinical chemical parameters in guinea pigs. Toxicol.
    appl. Pharmacol., 51: 329-339.

    GASIEWICZ, T.A. & RUCCI, G. (1984) Cytosolic receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Evidence for a homologous nature
    among various mammalian species. Mol. Pharmacol., 26: 90-98.

    GASIEWICZ, T.A., HOLSCHER, M.A., & NEAL, R.A. (1980) The effect of
    total parenteral nutrition on the toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Toxicol. appl.
    Pharmacol., 54: 469-488.

    GASIEWCICZ, T.A., OLSON, J.R., GEIGER, L.H., & NEAL, R.A. (1983a)
    Absorption, distribution and metabolism of
    2,3,7,8-tetrachlorodibenzodioxin (TCDD) in experimental animals.
    Environ. Sci. Res., 26: 495-525.


    GASIEWICZ, T.A., GEIGER, L.E., RUCCI, G., & NEAL, R.A. (1983b)
    Distribution, excretion, and metabolism of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin in C57Bl/6J, DBA/2J, and B602F,/J
    mice. Drug Metab. Disp., 11: 397-403.

    GASIEWICZ, T.A., NESS, W.C., & RUCCI, G. (1984) Ontogeny of the
    cytosolic receptor for 2,3,7,8-tetrachlorodibenzo-p-dioxin in rat
    liver, lung, and thymus. Biochem. biophys. Res. Commun., 118:
    183-190.

    GASIEWICZ, T.A., RUCCI, G., HENRY, E.C., & BAGGS, R.B. (1986) Changes
    in hamster hepatic cytochrome P-450, ethoxycoumarin o-deethylase, and
    reduced NAD(P): Menadione oxidoreductase following treatment with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem. Pharmacol., 35:
    2737-2742.

    GEBEFUGI, I., BAUMANN, R., & KORTE, F. (1977) [Photochemical breakdown
    of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) under stimulated
    environmental conditions.] Naturwissenschaften, 64: 486-487 (in
    German).

    GEIGER, L.E. & NEAL, R.A. (1981) Mutagenicity testing of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in histidine auxotrophs of
    Salmonella typhimurium. Toxicol. appl. Pharmacol., 59: 125-129.

    GERMANY (1985) [Review of dioxins], Berlin, Erich Schmidt Verlag, pp.
    257-266 (Federal Office for the Environment, Report 5/85, November
    1984) (in German).

    GIAVINI, E., PRATI, M., & VISMARA, C. (1982a) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin administered to pregnant rats
    during the pre-implantation period. Environ. Res., 29: 185-189.

    GIAVINI, E., PRATI, M., & VISMARA, C. (1982b) Rabbit teratology study
    with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ. Res., 27:
    74-78.

    GIAVINI, E., PRATI, M., & VISMARA, C. (1983) Embryotoxic effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin administered to female rats before
    mating. Environ. Res., 31: 105-110.

    GIERTHY, J.F. & CRANE, D. (1984) Reversible inhibition of in vitro
    epithelial cell proliferation by 2,3,7,8-tetrachloro-dibenzo-p-dioxin.
    Toxicol. appl. Pharmacol., 74: 91-98.

    GIERTHY, J.F. & CRANE, D. (1985a) In vitro bioassay for dioxin-like
    activity based on alterations in epithelial cell proliferation and
    morphology. Fundam. appl. Toxicol., 5: 754-759.


    GIERTHY, J.F. & CRANE, D. (1985b) Development of in vitro
    bioassays for chlorinated dioxins and dibenzofurans. In: Keith, L.H.,
    Rappe, C., & Choudhary, G., ed. Chlorinated dioxins and
    dibenzofurans in the total environment II, Stoneham, Maine,
    Butterworth Publishers, pp. 267-284.

    GIERTHY, J.F., CRANE, D., & FRENKEL, G.D. (1984) Application of an
    in vitro keratinization assay to extracts of soot from a fire in
    a polychlorinated biphenyl-containing transformer. Fundam. appl.
    Toxicol., 4: 1036-1041.

    GILBERT, P., SAINT-RUF, G., PONCELET, F., & MERCIER, M. (1980) Genetic
    effects of chlorinated anilines and azobenzenes on salmonella
    typhimurium. Arch. environ. Contam. Toxicol., 9: 533-541.

    GOLDMANN, P.J. (1972) [Very severe acute chloracne caused by
    trichlorophenol decomposition products.] Arbeitsmed. Sozialmed.
    Arbeitshyg., 7: 12-18 (in German).

    GOLDMANN, P.J. (1973) [Very severe acute chloracne: mass poisoning by
    2,3,6,7-tetrachlorodibenzodioxin.] Hautarzt, 24: 149-152 (in
    German).

    GOLDSTEIN, J.A. & LINKO, P. (1984) Differential induction of two
    2,3,7,8-tetrachlorodibenzo-p-dioxin-inducible forms of cytochrome
    P-450 in extrahepatic versus hepatic tissues. Mol. Pharmacol., 25:
    185-191.

    GOLDSTEIN, J.A., HICKMAN, P., BERGMAN, H., & VOS, J.G. (1973) Hepatic
    porphyria induced by 2,3,7,8-tetrachloro-dibenzo-p-dioxin in the
    mouse. Res. Commun. chem. Pathol. Pharmacol., 6: 919-928.

    GOLDSTEIN, J.A., MCKINNEY, J.D., LUCIER, G.W., HICKMAN, P., BERGMAN,
    H., & MOORE, J.A. (1976) Toxicological assessment of
    hexachlorobiphenyl isomers and 2,3,7,8-tetrachlorodibenzofuran in
    chicks. II. Effects on drug metabolism and porphyrin accumulation.
    Toxicol. appl. Pharmacol., 36: 81-92.

    GOLDSTEIN, J.A., HASS, J.R., LINKO, P., & HARVAN, D.J. (1978)
    2,3,7,8-Tetrachlorodibenzofuran in a commercially available 99% pure
    polychlorinated biphenyl isomer identified as the inducer of hepatic
    cytochrome P-448 and aryl hydrocarbon hydroxylase in the rat. Drug
    Metab. Disp., 6: 258-264.

    GOLDSTEIN, J.A., LINKO, P., & BERGMAN, H. (1982) Induction of
    porphyria in the rat by chronic versus acute exposure to
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem. Pharmacol., 31:
    1607-1613.


    GONZALES, F.J., TUKEY, R.H., & NEBERT, D.W. (1984) Structural gene
    products of the Ah locus. Transcriptional regulation of cytochrome
    P1-450 and P3-450mRNA levels by 3-methylchol-anthrene. Mol.
    Pharmacol., 26: 117-121.

    GORSKI, T., KONOPKA, L., & BRODZKI, M. (1984) Persistence of some
    polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans of
    pentachlorophenol in human adipose tissue. Rocz. Pzh, XXXV(4):
    297-301.

    GOTHE, R. & WACHTMEISTER, C.A. (1972) Synthesis of
    1,2,4,5,7,8-hexa-chloroxanthene. Acta chem. Scand., 26: 2523-2576.

    GOTZ, R. (1986) Chlorinated dioxins and dibenzofurans in leachate and
    sediments of the sanitary landfill in Hamburg-Georswerder.
    Chemosphere, 15: 1981-1984.

    GRAY, A.P., STEVEN, P.C., & CANTRELL, J.S. (1975) Intervention of the
    Smiles rearrangement in synthesis of dibenzo-p-dioxins:
    1,2,3,6,7,8-and 1,2,3,7,8,9-hexachlorodibenzo-dioxin. Tetrahydron
    Lett., 33: 2873-2876.

    GRAY, A.P., STEVEN, P.C., SOLOMON, I.J., & ANILINE, O. (1976)
    Synthesis of specific polychlorinated dibenzo-p-dioxins. J. org.
    Chem., 41: 2435-2437.

    GREEN, S. & MORELAND, F.S. (1975) Cytogenetic evaluation of several
    dioxins in the rat. Toxicol. appl. Pharmacol., 33: 161.

    GREEN, S., MORELAND, F., & SHEU, C. (1977) Cytogenetic effect of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on rat bone marrow cells. FDA
    By-lines, 6: 292-294.

    GREENBURG, L., MAYERS, M.R., & SMITH, A.R. (1939) The systemic effects
    resulting from exposure to certain chlorinated hydro-carbons. J. ind.
    Hyg. Toxicol., 21: 29-38.

    GREENLEE, W.F. & POLAND, A. (1978) An improved assay of
    7-ethoxycoumarin O-deethylase activity: induction of hepatic enzyme
    activity in C57BL/6J and DBA/2J mice by phenobarbital,
    3-methylcholanthrene and 2,3,7,8-tetrachlorodibenzo-p-dioxin. J.
    Pharmacol. exp. Ther., 205: 596-605.

    GREENLEE, W.F. & POLAND, A. (1979) Nuclear uptake of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in C57BL/6J and DBA/2J mice. Role
    of the hepatic cytosol receptor protein. J. biol. Chem., 254:
    9814-9821.


    GREENLEE, W.F., DOLD, K.M., IRONS, R.D., & OSBORNE, R. (1985) Evidence
    for direct action of 2,3,7,8-tetrachlorodidibenzo-p-dioxin (TCDD) on
    thymic epithelium. Toxicol. appl. Pharmacol., 79: 112-120.

    GREIG, J.B. (1972) Effect of 2,3,7,8-tetrachlorodibenzo-1, 4-dioxin on
    drug metabolism in the rat. Biochem. Pharmacol., 21: 3196-3198.

    GREIG, J.B. & OSBORNE, G. (1981) Biochemical and morphological changes
    induced by 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat liver cell
    plasma membrane. J. appl. Toxicol., 1(6): 334-338.

    GREIG, J.B., JONES, G., BUTLER, W.H., & BARNES, J.M. (1973) Toxic
    effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Food Cosmet.
    Toxicol., 11: 585-595.

    GREIG, J.B., TAYLOR, D.M., & JONES, J.D. (1974) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on stimulated DNA synthesis in the
    liver and kidney of the rat. Chem.-biol. Interact., 8: 31-39.

    GREIG, J.B., FRANCIS, J.E., KAY, S.J.E., LOVELL, D.P., & SMITH, A.G.
    (1984) Incomplete correlation of 2,3,7,8-tetra-chlorodibenzo-p-dioxin
    hepatotoxicity with Ah phenotype in mice. Toxicol. appl. Pharmacol.,
    74: 17-25.

    GROSS, M.L., LAY, J.O., Jr, LYON, P.A., LIPPSTREU, D., KANGAS, N.,
    HARLESS, R.L., TAYLOR, S.E., & DUPUY, A.E. Jr (1984)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin levels in adipose tissue of
    Vietnam veterans. Environ. Res., 33: 261-268.

    GUENTHNER, T.M., FYSH, J.M., & NEBERT, D.W. (1979)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin: covalent binding of reactive
    metabolic intermediates principally to protein in vitro.
    Pharmacology, 19: 12-22.

    GUPTA, B.N., VOS, J.G., MOORE, J.A., ZINKL, J.G., & BULLOCK, B.C.
    (1973) Pathologic effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin in
    laboratory animals. Environ. Health Perspect., 5: 125-140.

    GURTOO, H.L., PARKER, N.B., PAIGEN, B., HAVENS, M.B., MINOWADA, J., &
    FREEDMAN, H.J. (1979) Induction, inhibition and some enzymological
    properties of aryl hydrocarbon hydroxylase in fresh mitogen-activated
    human lymphocytes. Cancer Res., 39: 4620-4629.

    GUSTAFSSON, J.-A. & INGELMAN-SUNDBERG, M. (1979) Changes in steroid
    hormone metabolism in rat liver microsomes following administration of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).  Biochem. Pharmacol.,
    28: 497-499.

    HAAPARANTA, T., GLAUMANN, H., & GUSTAFSSON, J.-A. (1983) Induction of
    cytochrome P-450 dependent reactions in the rat ventral prostate by
    beta-naphthoflavone and 2,3,7,8-tetra-chlorodibenzo-p-dioxin.
    Toxicology, 29: 61-75.

    HAGENMAIER, H. & BRUNNER, H. (1987) Isomer specific analysis of PCDD
    and PCDF in pentachlorophenol and sodium pentachloro- phenate samples
    in the sub-PPB range. Chemosphere, 16: 1759-1764.

    HAGENMAIER, H., DRAFT, M., JAGER, W., MAYER, U., LUETZKE, K., &
    SIEGAL, D. (1986) Comparison of various sampling methods for PCDDs and
    PCDFs in stack gas. Chemosphere, 15(9-12): 1187-1192.

    HAKANSSON, H. (1988) Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin
    of the fate of vitamin A in rodents, Stockholm, Department of
    Toxicology and Institute of Environmental Medicine, Karolinska
    Institute (Ph.D. Thesis).

    HAKANSSON, H. & AHLBORG, U.G. (1985) The effect of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the uptake, distribution
    and excretion of a single oral dose of 11,12, 3H-retinylacetate and on
    the vitamin A status in the rat. J. Nutr., 115: 759-771.

    HAKANSSON, H., AHLBORG, U.G., & GOTTLING, L. (1986) The effect of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the distribution and
    excretion of the endogenous pool of vitamin A in rats with low liver
    vitamin A stores. Chemosphere, 15: 1715-1723.

    HAKANSSON, H., WAERN, F., & AHLBORG, U.G. (1987) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the lactating rat on
    maternal and neonatal vitamin A status. J. Nutr., 117: 580-586.

    HANNAH, R.R., LUND, J., POELLINGER, L., GILLNER, M., & GUSTAFSSON,
    J.-A. (1986) Characterization of the DNA-binding properties of the
    receptor for 2,3,7,8-tetrachlorodibenzo-p-dioxin. Eur. J.
    Biochem., 156: 237-242.

    HARDELL, L. (1981) Relation of soft-tissue sarcoma, malignant lymphoma
    and colon cancer to phenoxy acids, chlorophenols and other agents.
    Scand. J. Work Environ. Health, 7: 119-130.

    HARDELL, L. & ERIKSSON, M. (1981) Soft-tissue sarcomas, phenoxy
    herbicides, and chlorinated phenols. Lancet, 2: 250.

    HARDELL, L. & SANDSTROM, A. (1979) Case-control study: soft-tissue
    sarcomas and exposure to phenoxyacetic acids or chlorophenols. Br.
    J. Cancer, 39: 711-717.

    HARDELL, L., ERIKSSON, M., LENNER, P., & LUNDGREN, E. (1981) Malignant
    lymphoma and exposure to chemicals, especially organic solvents,
    chlorophenols and phenoxy acids: a case-control study. Br. J.
    Cancer, 43: 169-176.

    HARLESS, R.L. & LEWIS, R.G. (1982) Quantitative determination of
    2,3,7,8-tetrachlorodibenzo-p-dioxin residues by gas
    chromatography/mass spectrometry. In: Hutzinger, O., ed. Chlorinated
    dioxins and related compounds, Oxford, London, Pergamon Press, pp.
    25-36.

    HARLESS, R.L., OSWALD, E.O., LEWIS, R.G., DUPUY, A.E., MCDANIEL, D.D.,
    & TAI, H. (1982) Determination of 2,3,7,8-tetrachlorodibenzo-p-dioxin
    in fresh water fish. Chemosphere, 11, 193-198.

    HARLESS, R.L., LEWIS, R.G., DUPUY, A.E., & MCDANIEL, D.D. (1983)
    Analyses for 2,3,7,8-tetrachlorodibenzo-p-dioxin residues in
    environmental samples. In: Tucker, R.E., ed. Human and
    environmental risks of chlorinated dioxins and related compounds,
    New York, London, Plenum Press, pp. 161-172.

    HARRIS, M.W., MOORE, J.A., VOS, J.G., & GUPTA, B.N. (1973) General
    biological effects of TCDD in laboratory animals. Environ. Health
    Perspect., 5: 101-109.

    HASSAN, M.Q., STOHS, S.J., & MURRAY, W.J. (1983) Comparative ability
    of TCDD to induce lipid peroxidation in rats, guinea pigs and syrian
    golden hamsters. Bull. environ. Contam. Toxicol., 31: 649-657.

    HASSAN, M.Q., STOHS, S.J., & MURRAY, W.J. (1985a) Inhibition of
    TCDD-induced lipid peroxidation, glutathione peroxidase activity and
    toxicity by BHA and glutathione. Bull. environ. Contam. Toxicol.,
    34: 787-796.

    HASSAN, M.Q., STOHS, S.J., & MURRAY, W.J. (1985b) Effects of vitamins
    E and A on 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-induced lipid
    peroxidation and other biochemical changes in the rat. Arch. environ.
    Contam. Toxicol., 14: 437-442.

    HASSAN, M.Q., STOHS, S.J., MURRAY, W.J., & BIRT,D.F. (1985c) Dietary
    selenium, glutathione peroxidase activity, and toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. J. Toxicol. environ. Health,
    15: 405-415.

    HASSOUN, E.M. & DENCKER, L. (1982) TCDD embryotoxicity in the mouse
    may be enhanced by beta-naphtoflavone, another ligand of the
    Ah-receptor. Toxicol. Lett., 12: 191-198.

    HASSOUN, E.M., D'ARGY, R., & DENCKER, L. (1984a) Teratogenicity of
    2,3,7,8-tetrachlorodibenzofuran in the mouse. J. Toxicol. environ.
    Health, 14: 337-351.

    HASSOUN, E.M., D'ARGY, R., DENCKER, L., LUNDING, L.G., & BORWELL, P.
    (1984b) Teratogenicity of 2,3,7,8-tetrachloro-dibenzofuran in BXD
    recombinant inbred strains. Toxicol. Lett., 23: 37-42.

    HAY, A. (1984) Experimental toxicology and cytogenetics: an overview.
    In: Westing, A.N., ed. Herbicides in war, the long-term ecological
    and human consequences, London, Taylor & Francis, pp. 161-166.

    HAY, A.W.M. (1977) Tetrachlorodibenzo-p-dioxin release at Seveso.
    Disasters, 1: 289-308.

    HAYABUCHI, H., YOSHIMURA, T., & KURATSUNE, M. (1979) Consumption of
    toxic rice oil by Yusho patients and its relation to the clinical
    response and latent period. Food Cosmet. Toxicol., 17: 455-461.

    HAYES, K.C. (1971) On the pathophysiology of vitamin a deficiency.
    Nutr. Rev., 29: 3-6.

    HELLING, C.S., ISENSEE, A.R., WOOLSON, E.A., ENZOR, P.D.J., JONES,
    G.E., PLIMMER, J.R., & KEARNEY, P.C. (1973) Chlorodioxins in
    pesticides, soils and plants. J. environ. Qual., 2: 171-178.

    HENCK, J.M., NEW, M.A., KOBICA, R.J., & RAO, K.S. (1981)
    2,3,7,8-tetrachlorodibenzo-p-dioxin: acute oral toxicity in hamsters.
    Toxicol. appl. Pharmacol., 59: 405-407.

    HERXHEIMER, K. (1899) [Chloracne.] Mnchenere med. Wochenschr.,
    46: 278 (in German).

    HERZBERG, J.J. (1947) [Chloracne after ingestion of chlorinated
    paraffin.] Dermatol. Wochenschr., 7: 425-433 (in German).

    HINSDILL, R.D., COUCH, D.L., & SPEIRS, R.S. (1980) Immunosuppression
    in mice induced by dioxin (TCDD) in feed. J. environ. Pathol.
    Toxicol., 3: 401-425.

    HIROSAWA, K. & YAMADA, E. (1973). The localization of the vitamin A in
    the mouse liver as revealed by electron microscope radioautography.
    J. electron. Microsc., 22: 337-346.

    HOFFMAN, R.E., STEHR-GREEN, P.A., WEBB, K.B., EVANS, R.G., KNUTSEN,
    A.P., SCHRAMM, W.F., STAAKE, J.L., GIBSON, B.B., & STEINBERG, K.K.
    (1986) Health effects of long-term exposure to
    2,3,7,8-tetrachlorodibenzo-p-dioxin. J. Am. Med. Assoc., 255:
    2031-2038.

    HOFMANN, H.Th. (1957) [Recent findings with highly toxic chlorinated
    hydrocarbons.] Arch. exp. Pathol. Pharmakol., 232: 228-230 (in
    German).

    HOFMANN, M.F. & MENEGHINI, C.L. (1962) [Folliculosis from
    chlorine-substituted hydrocarbons (chloracne).] G. Ital. Dermatol.,
    103: 427-450 (in Italian).

    HOLMSTEDT, B. (1980) Prolegomena to Seveso. Arch. Toxicol., 44:
    211-230.

    HONCHAR, P.A. & HALPERIN, W.E. (1981) 2,4,5-trichlorophenol and
    soft-tissue sarcoma. Lancet, 21: 268-269.

    HOOK, G.E.R., HASEMAN, J.K., & LUCIER, G.W. (1975a) Induction and
    suppression of hepatic and extrahepatic microsomal
    foreign-compound-metabolizing enzyme systems by
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Chem.-biol. Interact., 10:
    199-214.

    HOOK, G.E.R., ORTON, T.C., MOORE, J.A., & LUCIER, G.W. (1975b)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin-induced changes in the
    hydroxylation of biphenyl by rat liver microsomes. Biochem.
    Pharmacol., 24: 335-340.

    HORI, S., OBANA, H., TANAKA, R., & KASHIMOTO, T. (1986) Comparative
    toxicity in rats of polychlorinated biphenyls (PCBs), polychlorinated
    quaterphenyls (PCQs) and poly-chlorinated dibenzofurans (PCDFs)
    present in rice oil causing "Yusho". Eisei Kagaku, 32: 13-21.

    HRYHORCZUK, D.O., WITHROW, W.A., HESSE, C.S., & BEASLEY, V.R. (1981)
    A wire reclamation incinerator as a source of environmental
    contamination with tetrachlorodibenzo-p-dioxins and
    tetrachlorodibenzofurans. Arch. environ. Health, 36: 228-234.

    HSU, S., MA, C., HSU, S.K., WU, S., HSU, N.H., & YEH, C. (1984)
    Discovery and epidemiology of PCB poisoning in Taiwan. Am. J. ind.
    Med., 5: 71-79.

    HUDSON, L.G., SHAIKH, R., TOSCANO, W.A., Jr, & GREENLEE, W.F. (1983)
    Induction of 7-ethoxycoumarin o-deethylase activity in cultured human
    epithelial cells by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD):
    Evidence for TCDD receptor. Biochem. biophys. Res. Commun., 115:
    611-617.

    HUDSON, L.G., TOSCANO, W.A., Jr, & GREENLEE, W.F. (1985) Regulation of
    epidermal growth factor binding in a human keratinocyte cell line by
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    77: 251-259.

    HUDSON, L.G., TOSCANO, W.A., Jr, & GREENLEE, W.F. (1986)
    2,3,7,8-Tetrachorodibenzo-p-dioxin (TCDD) modulates epidermal growth
    factor (EGF) binding to basal cells from a human keratinocyte cell
    line. Toxicol. appl. Pharmacol., 82: 481-492.

    HUFF, J.E., MOORE, J.A., SARACCI, R., & TOMATIS, L. (1980) Long-term
    hazards of polychlorinated dibenzodioxins and polychlorinated
    dibenzofurans. Environ. Health Perspect., 36: 221-240.

    HUQUE, T. (1981) Excretion of radioactive metabolites of retinal as an
    index of vitamin A status in rats. Nutr. Rep. Int., 24: 171-179.

    HUSSAIN, S., EHRENBERG, L., LOFROTH, G., & GEJVALL, T. (1972)
    Mutagenic effects of TCDD on bacterial systems. Ambio, 1: 32-33.

    HUTZINGER, O., SAFE, S., WENTZELL, B.R., & ZITKO, V. (1973)
    Photochemical degradation of di-and octachlorodibenzofuran. Environ
    Health Perspect., 5: 267-271.

    HWANG, S.W. (1973) Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on
    the biliary excretion of indocyanine Green in rat. Environ. Health
    Perspect., 5: 227-231.

    IARC (1977) Some fumigants, the herbicides 2,4-D and 2,4,5-T,
    chlorinated dibenzodioxins and miscellaneous industrial chemicals,
    Lyon, France, International Agency for Research on Cancer, pp. 41-102
    (IARC Monographs on the Evaluation of the Carcinogenic Risk of
    Chemicals to Humans, Vol. 15).

    IARC (1982) Chemicals, industrial processes and industries
    associated with cancer in humans, Lyon, France, International Agency
    for Research on Cancer (IARC Monograph on the Evaluation of the
    Carcinogenic Risk of Chemicals to Humans, Suppl. No. 4).

    IARC (1986) Some halogenated hydrocarbons and pesticide exposures,
    Lyon, International Agency for Research on Cancer (IARC Monographs on
    the Evaluation of the Carcinogenic Risk of Chemicals to Humans, Vol.
    41).

    IDEO, G., BELLATI, G., BELLOBUONO, A., MOCARELLI, P., MAROCCHI, A., &
    BRAMBILLA, P. (1982) Increased urinary D-glucaric acid excretion by
    children living in an area polluted with
    tetrachlorodibenzo-para-dioxin (TCDD). Clin. Chim. Acta, 120:
    273-283.

    IDEO, G., BELLATI, G., BELLOBUONO, A., & BISSANTI, L. (1985) Urinary
    d-glucaric acid excretion in the Seveso area, polluted by
    tetrachlorodibenzo-p-dioxin (TCDD): Five years of experience.
    Environ. Health Perspect., 60: 151-157.

    INNAMI, S.I., NAKAMURA, A., MIYAZAKI, M., NAGAYAMA, S., & NISHIDE, E.
    (1976) Further studies on the reduction of vitamin A content in the
    liver of rats given polychlorinated biphenyls. J. nutr. Sci.
    Vitaminol., 22: 409-418.

    IOANNOU, Y.M., BIRNBAUM, L.S., & MATTHEWS, H.B. (1983) Toxicity and
    distribution of 2,3,7,8-tetrachlorodibenzofuran in male guinea pigs.
    J. Toxicol. environ. Health, 12: 541-553.

    IRPTC (1987) IRPTC Legal file 1986, Geneva, International Register
    of Potentially Toxic Chemicals, United Nations Environment Programme,
    Geneva, Switzerland, Vol. I (Data Profile Series No. 7).

    ISENSEE, A.R. (1978) Bioaccumulation of
    2,3,7,8-tetrachloro-dibenzo-para-dioxin. Ecol. Bull. (Stockholm),
    27: 255-262.

    ISENSEE, A.R. & JONES, G.E. (1971) Absorption and trans-location of
    root and foliage applied 2,4-dichlorophenol,
    2,7-dichlorodibenzo-p-dioxin and 2,3,7,8-tetrachlorodibenzo-p-dioxin.
    J. agric. food. Chem., 19: 1210-1214.

    ISENSEE, A.R. & JONES, G.E. (1975) Distribution of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in aquatic model ecosystem.
    Environ. Sci. Technol., 9: 668-672.


    ISRAEL, D.I. & WHITLOCK, J.P., Jr (1983) Induction of mRNA specific
    for cytochrome P1-450 in wild type and variant mouse hepatoma cells.
    J. biol. Chem., 258: 10390-10394.

    ISRAEL, D.I. & WHITLOCK, J.P., Jr (1984) Regulation of cytochrome
    P1-450 gene transcription by 2,3,7,8-tetrachloro-dibenzo-p-dioxin in
    wild type and variant mouse hepatoma cells. J. biol. Chem., 259:
    5400-5402.

    JANSING, R.L. & SHAIN, W. (1985) Aryl hydrocarbon hydroxylase
    induction in adult rat hepatocytes in primary culture by several
    chlorinated aromatic hydrocarbons including
    2,3,7,8-tetrachloridibenzo-p-dioxin. Fundam. appl. Toxicol., 5:
    713-720.

    JENSEN, D.J. & HUMMEL, R.A. (1982) Secretion of TCDD in milk and cream
    following the feeding of TCDD to lactating dairy cows. Bull.
    environ. Contam. Toxicol., 29: 440-446.

    JENSEN, D.J., GETZENDANER, M.E., HUMMEL, R.A., & TURLEY, J. (1983)
    Residue studies for (2,4,5-trichlorophenoxy)acetic acid and
    2,3,7,8-tetrachlorodibenzo-p-dioxin in grass and rice. J. agric.
    food Chem., 31: 118-122.

    JENSEN, N.E. & WALKER, A.E. (1972) Chloracne: Three cases. Proc. R.
    Soc. Med., 65: 687-688.

    JIRASEK, L., KALENSKY, J., & KUBEC, K. (1973) [Acne chlorine and
    porphyria-cutanea tarda by the production of herbicides.] Cs.
    dermatol., 48: 306-317 (in Czech).

    JIRASEK, L., KALENSKY, J., KUBEC, K., PAZDEROVA, J., & LUKAS, E.
    (1976) [Chloracne, porphyria-cutanea tarda and other intoxication by
    herbicides.] Hautarzt, 27: 328-333 (in German).

    JOHANSSON, G., GILLNER, M., HOGBERG, B., & GUSTAFSSON, J.-A. (1982)
    The TCDD receptor in rat intestinal mucosa and its possible dietary
    ligands. Nutr. Cancer, 3: 134-143.

    JOHNSON, E.F. & MULLER-EBERHARD, U. (1977a) Purification of the major
    cytochrome P-450 of liver microsomes from rabbits treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Biochem. biophys. Res.
    Commun., 76: 652-659.

    JOHNSON, E.F. & MULLER-EBERHARD, U. (1977b) Resolution of two forms of
    cytochrome P-450 from liver microsomes of rabbits treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. J. biol. Chem., 252:
    2839-2845.


    JOHNSON, E.F., SCHWAB, G.E., & MULLER-EBERHARD, U. (1979) Multiple
    forms of cytochrome P-450: Catalytic differences exhibited by two
    homogeneous forms of rabbit cytochrome P-450. Mol. Pharmacol., 15:
    708-718.

    JOHNSON, F.E., KUGLER, N.A., & BROWN, S.M. (1981) Soft tissue sarcomas
    and chlorinated phenols. Lancet, 1: 40.

    JONES, D., SAFE, S., MORCOM, E., HOLCOMB, C., COPPOCK, C., & IVIE, W.
    (1987) Bioavailability of tritiates
    2,3,7,8-tetra-chlorodibenzo-p-dioxin administered to Holstein dairy
    cows. Chemosphere, 16: 1743-1748.

    JONES, E.L. & KRIZEK, H. (1962) A technic for testing acnegenic
    potency in rabbits, applied to the potent acnegen,
    2,3,7,8-tetrachlorodibenzo-p-dioxin. J. invest. Dermatol., 39:
    511-517.

    JONES, G. (1975) A histochemical study of the liver lesion induced by
    2,3,7,8-tetrachlorodibenzo-p-dioxin (dioxin) in rats. J. Pathol.,
    116: 101-105.

    JONES, G. & BUTLER, W.H. (1974) A morphological study of the liver
    lesion induced by 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats. J.
    Pathol., 112: 93-97.

    JONES, G. & GREIG, J.B. (1975) Pathological changes in the liver of
    mice given 2,3,7,8-tetrachlorodibenzo-p-dioxin. Experientia (Basel),
    31: 1315-1317.

    JONES, K.G. & SWEENEY, G.D. (1977) Association between induction of
    aryl hydrocarbon hydroxylase and depression of uroporphyrinogen
    decarboxylase activity. Res. Commun. chem. Pathol. Pharmacol., 17:
    631-637.

    JONES, K.G. & SWEENEY, G.D. (1980) Dependence of the porphyrogenic
    effect of 2,3,7,8-tetrachlorodibenzo(p)dioxin upon inheritance of aryl
    hydrocarbon hydroxylase responsiveness. Toxicol. appl. Pharmacol.,
    53: 42-49.

    JONES, P.A. (1981) Chlorophenols and their impurities in the
    Canadian environment, Ottawa, Environment Canada (EPS 3-EC-81-2).

    JONES, P.A. (1984) Chlorophenols and their impurities in the
    Canadian environment: 1983 supplement, Ottawa, Environment Canada
    (EPS 3-EP-84-3).

    JONES, P.B.C., MILLER, A.G., ISRAEL, D.I., GALEAZZI, D.R., & WHITLOCK,
    J.P., Jr (1984) Biochemical and genetic analysis of variant mouse
    hepatoma cells which overtranscribe the cytochrome P1-450 gene in
    response to 2,3,7,8-tetrachloro-dibenzo-p-dioxin. J. biol. Chem.,
    259: 12357-12363.

    JONES, P.B.C., GALEAZZI, D.R., & WHITLOCK, J.P., Jr (1985) Control of
    cytochrome P1-450 gene expression by dioxin. Science, 227:
    1499-1502.

    JONES, P.B.C., DURRIN, L.K., GALEAZZI, D.R., & WHITLOCK, J.P., Jr
    (1986) Control of cytochrome P1-450 gene expression: Analysis of a
    dioxin-responsive enhancer system. Proc. Natl Acad. Sci. (USA),
    83: 2802-2806.

    JONES, R.E. & CHELSKY, M. (1986) Further discussion concerning
    porphyria cutanea tarda and TCDD exposure. Arch. environ. Health,
    41: 100-103.

    JOSEPHSON, J. (1983) Chlorinated dioxins and furans in the
    environment. Environ. Sci. Technol., 17: 124A-128A.

    JUNK, G.A. & RICHARD, J. (1981) Dioxins not detected in effluents from
    coal/refuse combustion. Chemosphere, 10: 1237-1241.

    KARASEK, F.W. & ONUSKA, F.I. (1982) Trace analysis of the dioxins.
    Anal. Chem., 54: 309A-324A.

    KARENLAMPI, S.O., EISEN, H.J., HANKINSON, O., & NEBER, D.W. (1983)
    Effects of cytochrome Pl-450 inducers on the cell-surface receptors
    for epidermal growth factor, phorbol 12,13-dibutyrate, or insulin of
    cultured mouse hepatoma cells. J. biol. Chem., 17: 10378-10383.

    KEARNEY, P.C., WOOLSON, E.A., & ELLINGTON, C.P., Jr (1972) Persistence
    and metabolism of chlorodioxins in soils. Environ. Sci. Technol.,
    6: 1017-1019.

    KEESEY, R.E., BOYLE, P.C., KEMNITZ, J.W., & MITCHEL, J.S. (1976) The
    role of the lateral hypothalamus in determining the body weight set
    point. In: Novin, D., Wyrwicks, W., & Bray, G., ed. Hunger: Basic
    mechanisms and clinical implications, New York, Raven Press, pp.
    243-255.

    KELLING, C.K., CHRISTIAN, B.J., INHORN, S.L., & PETERSON, R.E. (1985)
    Hypophagia-induced weight loss in mice, rats, and guinea pigs treated
    with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Fundam. appl. Toxicol.,
    5: 700-712.

    KENDE, A.S., WADE, J.J., RIDGE, D., & POLAND, A. (1974) Synthesis and
    fourier transform carbon - 13 nuclear magnetic resonance spectroscopy
    of new toxic polyhalodibenzo-p-dioxins. J. org. Chem., 39:
    931-937.

    KERKVLIET, N.I., BRAUNER, J.A., & MATLOCK, J.P. (1985) Humoral
    immunotoxicity of polychlorinated diphenyl ethers, phenoxy-phenols,
    dioxins and furans present as contaminants of technical grade
    pentachlorophenol. Toxicology, 36: 307-324.

    KEYS, B., HLAVINKA, M., MASON, G., & SAFE, S. (1985) Modulation of rat
    hepatic microsomal testosterone hydroxylases by
    2,3,7,8-tetrachlorodibenzo-p-dioxin and related toxic isostereomers.
    Can. J. Pharmacol., 63: 1537-1542.

    KEYS, B., PISKORSKA-PLISZCZYNSKA, J., & SAFE, S. (1986)
    Polychlorinated dibenzofurans as 2,3,7,8-TCDD antagonists: in
    vitro inhibition of monooxygenase enzyme induction. Toxicol.
    Lett., 31: 151-158.

    KHERA, K.S. & RUDDICK, J.A. (1973) Polychlorodibenzo-p-dioxins:
    Perinatal effects and the dominant lethal test in Wistar rats. Adv.
    Chem. Ser., 120: 70-84.

    KIMBLE, B.J. & GROSS, M.L. (1980) Tetrachlorodibenzo-p-dioxin
    quantitation in stack-collected coal fly ash. Science, 207: 59-61.

    KIMBROUGH, R.D. (1974) The toxicity of polychlorinated polycyclic
    compounds and related chemicals. CRC Crit. Rev. Toxicol., 2:
    445-498.

    KIMBROUGH, R.D. (1979) The carcinogenic and other chronic effects of
    persistent halogenated organic compounds. Ann. N.Y. Acad. Sci.,
    320: 415-418.

    KIMBROUGH, R.D. (1984) The epidemiology and toxicology of TCDD. Bull.
    environ. Contam. Toxicol., 33: 636-647.

    KIMBROUGH, R.D., GAINES, T.B., & LINDER, R.E. (1974)
    2,4-Dichlorophenyl-p-nitrophenyl ether (TOK). Effects on the lung
    maturation of rat fetus. Arch. environ. Health, 28: 316-319.

    KIMBROUGH, R.D., CARTER, C.D., LIDDLE, J.A., CLINE, R.E., & PHILLIPS,
    P.E. (1977) Epidemiology and pathology of a tetrachlorodibenzodioxin
    poisoning episode. Arch. environ. Health, 32: 77-86.


    KIMBROUGH, R.D., FALK, H., & STEHR, P. (1984) Health implications of
    2,3,7,8-tetrachlorodibenzodioxin (TCDD) contamination of residential
    soil. J. Toxicol. environ. Health, 14: 47-93.

    KIMMIG, J. & SCHULZ, K.H. (1957a) [Occupational acne (so called
    chloracne) caused by chlorinated aromatic cyclic ethers.]
    Dermatologica, 115: 540-546 (in German).

    KIMMIG, J. & SCHULZ, K.H. (1957b) [Chlorinated aromatic cyclic ethers
    as cause of chloracne.] Naturwissenschaften, 11: 337-338 (in
    German).

    KIMURA, S., GONZALEZ, F.J., & NEBERT, D.W. (1986) Tissue-specific
    expression of the mouse dioxin-inducible P1450 and P3450 genes:
    Differential transcriptional activation and mRNA stability in liver
    and extrahepatic tissues. Mol. cell. Biol., 6: 1471-1477.

    KING, F.G., DEDRICK, R.L., & COLLINS, J.M. (1983) Physiological model
    for the pharmacokinetics of 2,3,7,8-tetrachloro-dibenzofuran in
    several species. Toxicol. appl. Pharmacol., 67: 390-400.

    KITCHIN, K.T. & WOODS, J.S. (1979) 2,3,7,8-Tetrachlorodibenzo-p-dioxin
    (TCDD) effects on hepatic microsomal cytochrome P-448-mediated enzyme
    activities. Toxicol. appl. Pharmacol., 47: 537-546.

    KLEU, G. & GOLTZ, R. (1971) [Late and lasting damage resulting from
    the chronic occupational action of chlorophenol compounds.] Med.
    Klin., 66: 53-58 (in German).

    KNUTSON, J.C. & POLAND, A. (1980a)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin: failure to demonstrate toxicity
    in twenty-three cultured cell types. Toxicol. appl. Pharmacol.,
    54: 377-383.

    KNUTSON, J.C. & POLAND, A. (1980b) Keratinization of mouse teratoma
    cell line XB produced by 2,3,7,8-tetrachlorodibenzo-p-dioxin: an in
    vitro model of toxicity. Cell, 22: 27-36.

    KNUTSON, J.C. & POLAND, A. (1982) Response of murine epidermis to
    2,3,7,8-tetrachlorodibenzo-p-dioxin: Interaction of the Ah and hr
    Loci. Cell, 30: 225-234.

    KNUTSON, J.C. & POLAND, A. (1984) 2,3,7,8-Tetrachlorodibenzo-p-dioxin:
    Examination of biochemical effects involved in the proliferation and
    differentiation of XB cells. J. cell. Physiol., 121: 143-151.

    KOCHER, C.W., MAHLE, N.H., HUMMEL, R.A., & SHADOFF, L.A. (1978) A
    search for the presence of 2,3,7,8-tetrachloro-dibenzo-p-dioxin in
    beef fat. Bull. environ. Contam. Toxicol., 19: 229-236.

    KOCHMAN, S., BERNARD, J., CAZABAT, A., LAVAUD, F., LORTON, C., &
    RAPPE, C. (1986) Phenotypical dissection of immunoregulatory T cell
    subsets in human after furan exposure. Chemosphere, 15: 1799-1804.

    KOCIBA, R.J., KEELER, P.A., PARK, C.N., & GEHRING, P.J. (1976)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD): Results of a 13-week oral
    toxicity study in rats. Toxicol. appl. Pharmacol., 35: 553-574.

    KOCIBA, R.J., KEYES, D.G., BEYER, J.E., CARREON, R.M., WADE, C.E.,
    DITTENBER, D., KALNINS, R., FRAUSON, L., PARK, C.N., BARNARD, S.,
    HUMMEL, R., & HUMISTON, C.G. (1978) Results of a two year chronic
    toxicity and oncogenicity study of 2,3,7,8-tetrachlorodibenzo-p-dioxin
    (TCDD) in rats. Toxicol. appl. Pharmacol., 46: 279-303.

    KOCIBA, R.J., KEYES, D.G., LISOWE, R.W., KALNINS, R.P., DITTENBER,
    D.D., WADE, C.E., GORZINSKI, S.J., HAHLE, N.H., & SCHWETZ, B.A.
    (1979a) Results of a two-year chronic toxicity and encogenic study of
    rats ingesting diets containing 2,4,5-trichlorophenoxyacetic acid
    (2,4,5-T). Food Cosmet. Toxicol., 17: 205-221.

    KOCIBA, R.J., KEYES, D.G., BEYER, J.E., CARREON, R.M., & GEHRING, P.J.
    (1979b) Long-term toxicologic studies of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in laboratory animals.
    Ann. N.Y. Acad. Sci., 320: 397-404.

    KOHLI, K.K. & GOLDSTEIN, J.A. (1981) Effects of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin on hepatic and renal
    prostaglandin synthetase. Life Sci., 29: 299-305.

    KONDOROSI, A., FEDORCSAK, I., SOLYMOSY, F., EHRENBERG, L., &
    OSTERMAN-GOLKAR, S. (1973) Inactivation of Q-beta RNA by
    electrophiles. Mutat. Res., 17: 149-161.

    KOSHAKJI, R.P., HARBISON, R.D., & BUSH, M.T. (1984) Studies on the
    metabolic fate of (14C)2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in
    the mouse. Toxicol. appl. Pharmacol., 73: 69-77.

    KOURI, R.E., RATRIE, H., III, ATLAS, C.A., NIWA, A., & NEBERT, D.W.
    (1974) Aryl hydrocarbon hydroxylase induction in human lymphocyte
    cultures by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Life Sci., 15:
    1585-1595.


    KOURI, R.E., RUDE, T.H., & JOGLEKAR, R. (1978)
    2,3,7,8-Tetra-chlorodibenzo-p-dioxin as cocarcinogen causing
    3-methylchol-antrene-initiated subcutaneous tumors in mice genetically
    "non responsive" at Ah locus. Cancer Res., 38: 2777-2783.

    KRAUSE, L., VON & BRASSOW, H. (1978) [Contamnestic study of chloracne
    cases from the year 1954/55.] Arbeitsmed. Socialmed.
    Prventivmed., 13: 19-21 (in German).

    KROWKE, R. (1986) Studies on distribution and embryotoxicity of
    different PCDD and PCDF in mice and marmosets. Chemosphere, 15:
    2011-2022.

    KUNITA, N., KASHIMATO, T., MIYATA, H., FUKUSHIMA, S., HARI, S., &
    OBANA, H. (1984) Causal agents of Yusho. Am. J. ind. Med., 5:
    45-58.

    KUNTZMAN, D., LAWRENCE, D., & CONNEY, A.H. (1965) Michaelis constants
    for the hydroxylation of steroid hormones and drugs by rat liver
    microsomes. Mol. Pharmacol., 1: 163-167.

    KURL, R.N. & VILLEE, C.A. (1985) A metabolite of riboflavin binds to
    the 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) receptor.
    Pharmacology, 30: 241-244.

    KURL, R.N., LUND, J., POELLINGER, L., & GUSTAFSSON, J-A. (1982)
    Differential effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on nuclear
    RNA polymerase activity in the rat liver and thymus. Biochem.
    Pharmacol., 31: 2459-2462.

    KURL, R.N., LORING, J.M., & VILLEE, C.A. (1985) Control of
    2,3,7,8-tetrachlorodibenzo-p-dioxin binding protein(s) in the hamster
    kidney. Pharmacology, 30: 245-254.

    KUROKI, H., MASUDA, Y., YOSHIHARA, S., & YOSHIMURA, H. (1980)
    Accumulation of polychlorinated dibenzofurans in the livers of monkeys
    and rats. Food Cosmet. Toxicol., 18: 387-392.

    KUROKI, H., HARAGUCHI, K., & MASUDA, Y. (1984) Synthesis of
    polychlorinated dibenzofuran isomers and their gas chromato-graphic
    profiles. Chemosphere, 13: 561-573.

    KUROKI, J., KOGA, N., & YOSHIMURA, H. (1986) High affinity of
    2,3,4,7,8-pentachloridibenzofuran to cytochrome P-450 in the hepatic
    microsomes of rats. Chemosphere, 15: 731-738.

    LAHANIATIS, E.S., PARLAR, H., & KORTE, F. (1977) [Contributions to
    ecological chemistry CXXXII. On the occurrence of chlorinated
    hydrocarbons in the flue dust of refuse incinerators.] Chemosphere,
    7: 11-16 (in German).


    LAMB, J.C., MARKS, T.A., GLADEN, B.C., ALLEN, J.W., & MOORE, J.A.
    (1981) Male fertility, sister chromatid exchange, and germ cell
    toxicity following exposure to mixtures of chlorinated phenoxy acids
    containing 2,3,7,8-tetrachloro-dibenzo-p-dioxin. J. Toxicol.
    environ. Health, 8: 825-834.

    LAMB, J.C., IV, HARRIS, M.W., MCKINNEY, J.D., & BIRNBAUM, L.S. (1986)
    Effects of thyroid hormones on the induction of cleft palate by
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in C57Bl/6N mice. Toxicol.
    appl. Pharmacol., 84: 115-124.

    LAMPARSKI, L.L. & NESTRICK, T.J. (1981) Synthesis and identification
    of the 10 hexachlorodibenzo-p-dioxin isomers by high performance
    liquid and packed column gas chromatography. Chemosphere, 10:
    3-18.

    LAMPARSKI, L.L., NESTRICK, T.J., & STEHL, R.H. (1979) Determination of
    part-per-trillion concentration of 2,3,7,8-tetrachlorodibenzo-p-dioxin
    in fish. Anal. Chem., 51: 1453-1458.

    LAMPARSKI, L.L., NESTRICK, T.J., FRAWLEY, N.N., HUMMEL, R.A., KOCKER,
    C.W., MAHLE, N.H., MCCOY, J.W., MILLER, D.L., PETERS, T.L., PILLIPICH,
    J.L., SMITH, W.E., & TOBEY, S.W. (1986) Perspectives of a large scale
    environmental survey for chlorinated dioxins: Water analyses.
    Chemosphere, 15: 1445-1452.

    LANGER, H.G., BRADEY, T.P., & BRIGGS, P.R. (1973) Formation of
    dibenzodioxins and other condensation products from chlorinated
    phenols and derivatives. Environ. Health Perspect., 5: 3-7.

    LATHROP, G.D., WOLFE, W.H., ALBANESE, R.A., & MOYANAHAN, P.M. (1984)
    An epidemiologic investigation of health effects in air force
    personnel following exposure to herbicides. Baseline morbidity
    study results, San Antonio, Texas, USAF School of Aerospace Medicine
    (EK), Aerospace Medical Division, Brooks Air Force Base.

    LEE, P. & SUZUKI, K. (1980) Induction of aryl hydrocarbon hydroxylase
    activity in the rat prostate glands by
    2,3,7,8-tetrachlorodibenzo-p-dioxin. J. Pharmacol. exp. Ther.,
    215: 601-605.

    LIEM, H.H., MULLER-EBERHARD, U., & JOHNSON, E.F. (1980) Differential
    induction by 2,3,7,8-tetrachlorodibenzo-p-dioxin of multiple forms of
    rabbit microsomal cytochrome P-450: Evidence for tissue specificity.
    Mol. Pharmacol., 18: 565-570.

    LIGON, W.V., Jr & MAY, R.J. (1986) Determination of selected
    chlorodibenzofurans and chlorodibenzodioxins using two-dimensional gas
    chromatography/mass spectrometry. Anal. Chem., 58: 558-561.

    LINDAHL, R., RAPPE, C., & BUSER, H.R. (1980) Formation of
    polychlorinated dibenzofurans (PCDFs) and polychlorinated
    dibenzo-p-dioxins (PCDDs) from the pyrolysis of poly-chlorinated
    diphenyl ethers. Chemosphere, 9: 351-361.

    LINDAHL, R., ROPER, M., & DIETRICH, R.A. (1978) Rat liver aldehyde
    dehydrogenase-immunochemical identity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin inducible normal liver and
    2-acetylaminofluorene inducible hepatoma isozymes. Biochem.
    Pharmacol., 27: 2463-2465.

    LOPRIENO, N., SBRANA, I., RUSCIANO, D., LASCIALFARI, D., & LARI, T.
    (1982a) In vivo cytogenetic studies on mice and rats exposed to
    2,3,7,8-tetrachlorodibenzo-p-dioxin. In: Hutzinger, O., Frei, R.W.,
    Merian, E., & Pocchiari, P., ed. Chlorinated dioxins and related
    compounds. Impact on the environment, Oxford, New York, Pergamon
    Press, pp. 419-428.

    LOPRIENO, N., SBRANA, I., RUSCIANO, D., LASCIALFARI, D., LARI, T.,
    STRETTI, G., & FREZZA, D. (1982b) In vitro and in vivo genotoxicity
    studies on 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). In: Report
    of the 5th Meeting of the Seveso International Scientific Advisory
    Committee, Milan, Regione Lombardia (Report No. 32).

    LOVATI, M.R., GALBUSSERA, M., FRANCESCHINI, G., WEBER, G., RESI, L.,
    TANGANELLI, P., & SIRTORI, C.R. (1984) Increased plasma and aortic
    triglycerides in rabbits after acute administration of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    75: 91-97.

    LU, C.-J.H., BAGGS, R.B., REDMOND, D., HENRY, E.C., SCHECTER, A., &
    GASIEWICZ, T.A. (1986) Toxicity and evidence for meta-bolic
    alterations in 2,3,7,8-tetrachlorodibenzo-p-treated guinea pigs fed by
    total parenteral nutrition. Toxicol. appl. Pharmacol., 84:
    439-453.

    LUCIER, G.W., MCDANIEL, O.S., HOOK, G.E.R., FOWLER, B.A., SONAWANE,
    B.R., & FAEDER, E. (1973) TCDD-induced changes in rat liver microsomal
    enzymes. Environ. Health Perspect., 5: 199-209.

    LUCIER, G.W., MCDANIEL, O.S., & HOOK, G.E.R. (1975a) Nature of the
    enhancement of hepatic uridine diphosphate gluburonyl-transferase
    activity by 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats. Biochem.
    Pharmacol., 24: 325-334.


    LUCIER, G.W., SONAWANE, B.R., MCDANIEL, O.S., & HOOK, G.E.R. (1975b)
    Postnatal stimulation of hepatic microsomal enzymes following
    administration of TCDD to pregnant rats. Chem.-biol. Interact.,
    11: 15-26.

    LUCIER, G.W., RUMBAUGH, R.C., MCCOY, Z., HASS, R., HARVAN, D., &
    ALBRO, P. (1986) Ingestion of soil contaminated with
    2,3,7,8-tetrachlorodibenzop-dioxin (TCDD) alters hepatic enzyme
    activities in rats. Fundam. appl. Toxicol., 6: 364-371.

    LUSTENHOUWER, J.W.A., OLIE, K., & HUTZINGER, O. (1980) Chlorinated
    dibenzo-p-dioxins and related compounds in incinerator effluents: A
    review of measurements and mechanisms of formation. Chemosphere,
    9: 501-522.

    LUSTER, M.I., CLARK, G., LAWSON, L.D., & FAITH, R.E. (1979a) Effects
    of brief in vitro exposure to 2,3,7,8-tetrachloro-dibenzo-p-dioxin
    (TCDD) on mouse lymphocytes. J. environ. Pathol. Toxicol., 2:
    965-977.

    LUSTER, M.I., FAITH, R.E., & LAWSON, L.D. (1979b) Effects of
    2,3,7,8-tetrachlorodibenzofuran (TCDF) on the immune system in guinea
    pigs. Drug chem. Toxicol., 2: 49-60.

    LUSTER, M.I., BOORMAN, G.A., DEAN, J.H., HARRIS, M.W., LUEBKE, R.W.,
    PADARATHSINGH, M.L., & MOORE, J.A. (1980) Examination of bone marrow,
    immunologic parameters and host susceptibility following pre- and
    postnatal exposure to 2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD).
    Int. J. Immunopharmacol., 2: 301-310.

    LUSTER, M.I., TUCKER, A.N., HONG, L., BOORMAN, G.A., & PATTERSON, R.
    (1984) In vivo and in vitro effects of TCDD on stem cell and
    B cell differentiation. In: Biological mechanisms of dioxin action,
    Cold Spring Harbor, New York, Cold Spring Harbor Laboratory, pp.
    411-417 (Banbury Report No. 18).

    LUSTER, M.I., HONG, L.H., BOORMAN, G.A., CLARK, G., HAYES, H.T.,
    GREENLEE, W.F., DOLD, K., & TUCKER, A.N. (1985) Acute myelotoxic
    responses in mice exposed to 2,3,7,8-tetrachloro-dibenzo-p-dioxin
    (TCDD). Toxicol. appl. Pharmacol., 81: 156-165.

    MCCONNELL, E.E. (1980) Acute and chronic toxicity, carcinogenesis,
    reproduction, teratogenesis and mutagenesis in animals. In: Kimbrough,
    R.D., ed. Halogenated biphenyls, perphenyls, naphthalenes,
    dibenzodioxins and related products, Amsterdam, Oxford, New York,
    Elsevier Science Publishers, pp. 109-150.

    MCCONNELL, E.E. & MOORE, J.A. (1979) Toxicopathology characteristis of
    the halogenated aromatics. Ann. N.Y. Acad. Sci., 320: 138-150.

    MCCONNELL, E.E., MOORE, J.A., & DALGARD, D.W. (1978a) Toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in Rhesus monkeys (Macaca
    mulatta) following a single oral dose. Toxicol. appl. Pharmacol.,
    43: 175-187.

    MCCONNELL, E.E., MOORE, J.A., HASEMAN, J.K., & HARRIS, M.W. (1978b)
    The comparative toxicity of chlorinated dibenzo-p-dioxins in mice and
    guinea pigs. Toxicol. appl. Pharmacol., 44: 335-356.

    MCCONNELL, E.E., LUCIER, G.W., RUMBAUGH, R.C., ALBRO, P.W., HARVAN,
    D.J., HASS, J.R., & HARRIS, M.W. (1984) Dioxin in soil:
    bioavailability after ingestion by rats and guinea pigs. Science,
    223: 1077-1079.

    MCCUNE, E.L., SAVAGE, J.E., & O'DELL, B.L. (1962) Hydro-pericardium
    and ascites in chicks fed a chlorinated hydro-carbon. Poult. Sci.,
    41: 295-299.

    MCKINNEY, J., ALBRO, P., LUSTER, M., CORBETT, B., SCHROEDER, J., &
    LAWSON, L. (1982) Development and reliability of a radioimmunoassay
    for 2,3,7,8-tetrachlorodibenzo-p-dioxin. In: Hutzinger, O., ed.
    Chlorinated dioxins and related compounds: Impact on the
    environment, Oxford, New York, Pergammon Press, pp. 67-77.

    MCKINNEY, J.D. (1978) Analysis of
    2,3,7,8-tetrachlorodibenzo-para-dioxin in environmental samples. Ecol.
    Bull. (Stockholm), 27: 53-66.

    MCKINNEY, J.D., CHAE, K., GUPTA, B.N., MOORE, J.A., & GOLDSTEIN, J.A.
    (1976) Toxicological assessment of hexachlorobiphenyl isomers and
    2,3,7,8-tetrachlorodibenzo- furan in chicks. I. Relationship of
    chemical parameters. Toxicol. appl. Pharmacol., 36: 65-80.

    MCKINNEY, J.D., FAWKES, J., JORDAN, S., CHAE, K., OATLEY, S., COLEMAN,
    R.E., & BRINER, W. (1985) 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
    as a potent and persistent thyroxine agonist: a mechanistic model for
    toxicity based on molecular reactivity. Environ. Health Perspect.,
    61: 41-53.

    MCLAUGHLIN, D.L. & PEARSON, R.G. (1984) Concentrations of PCDD and
    PCDF in soil from the vicinity of the SWARU incinerator, Stoney
    Creek, July, 1983, Toronto, Ontario, Air Resources Branch, Ontario
    Ministry of the Environment, 22 pp.

    MCNULTY, W.P. (1984) Fetotoxicity of
    2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD) for rhesus macaques
    (Macaca mulatta). Am. J. Primatol., 6: 41-47.

    MCNULTY, W.P. (1985) Toxicity and fetotoxicity of TCDD, TCDF and PCB
    isomers in rhesus macaques (Macaca mulatta). Environ. Health
    Perspect., 60: 77-78.

    MCNULTY, W.P., POMERANTZ, I., & FARRELL, T. (1981) Chronic toxicity of
    2,3,7,8-tetrachlorodibenzofuran for rhesus macaques. Food Cosmet.
    Toxicol., 19: 57-65.

    MCNULTY, W.P., POMERANTZ, L.H., & FARRELL, T.J. (1982a) Chronic
    toxicity of 2,3,7,8-tetrachlorodibenzofuran for rhesus macaques. In:
    Hutzinger, O., ed. Chlorinated dioxins and related compounds:
    Impact on the environment, Oxford, New York, Pergamon Press, pp.
    411-418.

    MCNULTY, W.P., NIELSEN-SMITH, K.A., & LAY, J.O., Jr (1982b)
    Persistence of TCDD in monkey adipose tissue. Food Cosmet.
    Toxicol., 20: 985-987.

    MADGE, D.S. (1977) Effects of trichlorophenoxyacetic acid and
    chlorodioxins on small intestinal function. Gen. Pharmacol., 8:
    319-324.

    MADHUKAR, B.V. & MATSUMURA, F. (1981) Differences in the nature of
    induction of mixed-function oxidase systems of the rat liver among
    phenobarbital, DDT, 3-methylcholanthrene, and TCDD. Toxicol. appl.
    Pharmacol., 61: 109-118.

    MADHUKAR, B.V., BREWSTER, D.W., & MATSUMURA, F. (1984) Effects of in
    vivo-administered 2,3,7,8-tetrachlorodibenzo-p-dioxin on receptor
    binding of epidermal growth factor in the hepatic plasma membrane of
    rat, guinea pig, mouse, and hamster. Proc. Natl Acad. Sci. (USA),
    81: 7407-7411.

    MANARA, L., COCCIA, P., & CROCI, T. (1982) Persistent tissue levels of
    TCDD in the mouse and their reduction as related to prevention of
    toxicity. Drug Metab. Rev., 13(3): 423-446.

    MANARA, L., COCCIA, P., & CROCI, T. (1984) Prevention of TCDD toxicity
    in laboratory rodents by addition of charcoal or cholic acids to chow.
    Food chem. Toxicol., 22(10): 815-818.

    MANIS, J. & APAP, R. (1979) Intestinal organic anion transport,
    glutathione transferase and aryl hydrocarbon hydroxylase activity:
    effect of dioxin. Life Sci., 24: 1373-1380.

    MANIS, J. & KIM, G. (1977) Induction of intestinal iron transport by
    2,3,7,8-tetrachlorodibenzo-p-dioxin on environmental pollutant and
    potent inducer of aryl hydrocarbon hydroxylase. Clin. Res., 25:
    468A.

    MANIS, J. & KIM, G. (1979) Introduction of iron transport by a potent
    inducer of aryl hydrocarbon hydroxylase,
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Arch. environ. Health,
    34(3): 141-145.

    MANTOVANI, A., VECCHI, A., LUINI, W., SIRONI, M., CANDIANI, G.P.,
    SPREAFICO, F., & GARATTINI, S. (1980) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on macrophage and natural killer
    cell-mediated cytotoxicity in mice. Biomedicine, 32: 200-204.

    MARKLUND, S., KJELLER, L.-O., HANSSON, M., TYSKLIND, M., RAPPE, C.,
    RYAN, C., COLLAZO, H., & DOUGHERTY, R. (1986) Determination of PCDDs
    and PCDFs in incineration samples and pyrolytic products. In: Rappe,
    C., Choudhary, G., & Keith, L., ed. Chlorinated dioxins and
    dibenzofurans in perspective, Chelsea, Michigan, Lewis Publishers,
    pp. 79-92.

    MARKLUND, S., RAPPE, C., TYSKLIND, M., & EGEBACK, K. (1987)
    Identification of polychlorinated dibenzofurans and dioxins in
    exhausts from cars run on unleaded gasoline. Chemosphere, 16:
    29-36.

    MARPLE, L., BRUNCK, R., & THROOP, L. (1986a) Water solubility of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ. Sci. Technol., 20:
    180-182.

    MARPLE, L., BERRIDGE, B., & THROOP, L. (1986b) Measurement of the
    water-octanol partition coefficient of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Environ. Sci. Technol., 20:
    397-399.

    MASON, G. & SAFE, S. (1986) Synthesis, biologic and toxic effects of
    the major 2,3,7,8-tetrachlorodibenzo-p-dioxin metabolites in the rat.
    Toxicology, 41: 153-159.

    MASON, G., SAWYER, T., KEYS, B., BANDIERA, S., ROMKES, M.,
    PISKORSKA-PLISZCZYNSKA, J., ZMUDZKA, B., & SAFE, S. (1985)
    Polychlorinated dibenzofurans (PCDFs): correlation between in vivo
    and in vitro structure-activity relationships. Toxicology, 37:
    1-12.

    MASON, G., FARRELL, K., KEYS, B., PISKORSKA-PLISZCZYNSKA, J., SAFE,
    L., & SAFE, S. (1986) Polychlorinated dibenzo-p-dioxins: Quantitative
    in vitro and in vivo structure-activity relationships.
    Toxicology, 41: 21-31.


    MASON, M.E. & OKEY, A.B. (1982) Cytosolic and nuclear binding of
    2,3,7,8-tetrachlorodibenzo-p-dioxin to the Ah receptor in
    extra-hepatic tissues of rats and mice. Eur. J. Biochem., 123:
    209-215.

    MASUDA, Y., KUROKI, H., HARAGUCHI, K., & NAGAYAMA, J. (1985) PCB and
    PCDF congeners in the blood and tissues of Yusho and Yu-Cheng
    patients. Environ. Health Perspect., 59: 53-58.

    MASUDA, Y., KUROKI, H., HARAGUCHI, K., & NAGAYAMA, J. (1986) PCDFs and
    related compounds in humans from Yusho and Yu-Cheng incidents.
    Chemosphere, 15: 1621-1628.

    MATSUMURA, F. & BENEZET, H.J. (1973) Studies on the bioaccumu-lation
    and microbial degradation of 2,3,7,8-tetrachloro-dibenzo-p-dioxin.
    Environ. Health Perspect., 5: 253-258.

    MATSUMURA, F., QUENSEN, J., & TSUSHIMOTO, G. (1983) Microbial
    degradation of TCDD in a model ecosystem. In: Tucker, R.E., Young,
    A.L., & Gray, A.P., ed. Human and environmental risks of
    chlorinated dioxins and related compounds, New York, London, Plenum
    Press, pp. 191-220.

    MATSUMURA, F., BREWSTER, D.W., MADHUKAR, B.V., & BOMBICK, D.W. (1984)
    Alteration of rat hepatic plasma membrane functions by
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Arch. environ. Contam.
    Toxicol., 13: 509-515.

    MATTISON, D.R. & THORGEIRSSON, S.S. (1978) Gonadal aryl hydrocarbon
    hydroxylase in rats and mice. Cancer Res., 38: 1368-1373.

    MAY, G. (1973) Chloracne from the accidental production of
    tetrachlorodibenzodioxin. Br. J. ind. Med., 30: 276-283.

    MAY, G. (1982) Tetrachlorodibenzodioxin: a survey of subjects ten
    years after exposure. Br. J. ind. Med., 39: 128-135.

    MAZER, T., HILEMAN, F.D., NOBLE, R.W., & BROOKS, J.J. (1983) Synthesis
    of the 38 tetrachlorodibenzofuran isomers and identification by
    capillary column gas chromatography/mass spectrometry. Anal. Chem.,
    55: 104-110.

    MEYNE, J., ALLISON, D.C., BOSE, K., JORDAN, S.W., RIDOLPHO, P.F., &
    SMITH, J. (1985) Hepatotoxic doses of dioxin do not damage mouse bone
    marrow chromosomes. Mutat. Res., 157: 63-69.

    MILES, W.F., SINGH, J., GURPRASAD, N.P., & MALIS, G.P. (1985) Isomer
    specific determination of hexachlorodioxins in technical
    pentachlorophenol (PCP) and its sodium salt. Chemosphere, 14(6/7):
    807-810.


    MILLER, A.G., ISRAEL, D., & WHITLOCK, J.P., Jr (1983) Bio-chemical and
    genetic analysis of variant mouse hepatoma cells defective in the
    induction of benzo(a)pyrene-metabolizing enzyme. J. biol. Chem.,
    258: 3523-3527.

    MILNES, M.H. (1971) Formation of 2,3,7,8-tetrachlorodibenzo-dioxin by
    thermal decomposition of sodium 2,4,5-trichloro-phentane. Nature
    (Lond.), 232: 395-396.

    MILSTONE, L.M. & LAVIGNE, J.F. (1984)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin induces hyperplasia in confluent
    cultures of human keratinocytes. J. invest. Dermatol., 82: 532-534.

    MITCHUM, R.K., MOLER, G.F., & KORFMACHER, W.A. (1980) Combined
    capillary gas chromatography/atmospheric pressure negative chemical
    ionization/mass spectrometry for the determination of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in tissue. Anal. Chem., 52:
    2278-2282.

    MITTLER, J.C., ERTEL, N.H., PENG, R.X., YANG, C.S., & KIERNAN, T.
    (1984) Changes in testosterone hydroxylase activity in rat testis
    following administration of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Ann.
    N.Y. Acad. Sci., 438: 645-648.

    MIVRA, H., OMORI, A., & SHIBUE, M. (1974) The effect of chlorophenols
    on the excretion of porphyrins in urine. Jpn. J. ind. Health,
    16: 575-577.

    MOLLER, M.E., GLOWINSKI, I.B., & THORGEIRSSON, S.S. (1984) The
    genotoxicity of aromatic amines in primary hepatocytes iso-lated from
    C57BL/6 and DBA/2 mice. Carcinogenesis, 5: 797-804.

    MOORE, J.A., GUPTA, B.N., ZINKL, J.G., & VOS, J.G. (1973) Postnatal
    effects of maternal exposure to 2,3,7,8-tetrachloro-dibenzo-p-dioxins
    (TCDD). Environ. Health Perspect., 5: 81-85.

    MOORE, J.A., GUPTA, B.N., & VOS, J.G. (1976) Toxicity of
    2,3,7,8-tetrachloro-dibenzofuran - preliminary results. In:
    Proceedings from the National Conference on Polychlorinated
    Biphenyls, Chicago, Illinois, 19-21 November, 1975, Research
    Triangle Park, North Carolina, Research Triangle Park Institute.

    MOORE, J.A., MCCONNELL, E.E., DALGARD, D.W., & HARRIS, M.W. (1979)
    Comparative toxicity of three halogenated dibenzofurans in guinea
    pigs, mice and rhesus monkeys. Ann. N.Y. Acad. Sci., 320: 151-163.


    MOORE, R.W. & PETERSON, R.E. (1985) Enhanced catabolism and
    elimination of androgens do not cause the androgenic de-ficiency in
    2,3,7,8-tetrachlorodibenzo-p-dioxin-treated rats. Fed. Proc., 44:
    518.

    MOORE, R.W., POTTER, C.L., THEOBALD, H.M., ROBINSON, J.A., & PETERSON,
    R.E. (1985) Androgenic deficiency in male rats treated with
    2,3,7,8-tetra-chlorodbenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    79: 99-111.

    MORITA, M. & OISHI, S. (1977) Clearance and tissue distribution of
    polychlorinated dibenzofurans in mice. Bull. environ. Contam.
    Toxicol., 18: 61-66.

    MORITA, M., NAKAGAWA, J., & RAPPE, C. (1978) Polychlorinated
    dibenzofuran (PCDF) formation from PCB mixture by heat and oxygen.
    Bull. environ. Contam. Toxicol., 19: 665-670.

    MORTELMANS, K., HAWORTH, S., SPECK, W., & ZEIGER, E. (1984)
    Mutagenicity testing of agent orange components and related chemicals.
    Toxicol. appl. Pharmarcol., 75: 137-146.

    MOSES, M. & SELIKOFF, I.J. (1981) Soft tissue sarcomas, phenoxy
    herbicides, and chlorinated phenols. Lancet, 1: 1370.

    MOSES, M., LILIS, R., CROW, K.D., THORNTON, J., FISCHBEIN, A.,
    ANDERSON, H.A., & SELIKOFF, I.J. (1984) Health status of workers with
    past exposure to 2,3,7,8-tetrachloridibenzo-p-dioxin in manufacture of
    2,4,5-trichlorophenoxyacetic acid: Comparison of findings with and
    without chloracne. Am. J. ind. Med., 5: 161-182

    MOTTURA, A., ZEI, G., NUZZO, F., CRIMAUDO, C., GIORGI, R., VENERONI,
    P., PAGGINI, P., MACARELLI, P., FRACCARO, M., NICOLETTI, B., &
    DECARLI, L. (1981) Evaluation of results of chromosome analyses on
    lymphocytes of TCDD exposed subjects after the Seveso accident.
    Mutat. Res., 85: 238-239.

    MUKITARI, H., BAARS, A.J., & BREIMER, D.D. (1981) Differences in
    inducibility of particulate and cytosolic rat liver glutathione
    S-transferase activities. Xenobiotica, 11: 367-371.

    MULCAHY, M.T. (1980) Correspondence. Chromosome aberrations and "agent
    orange". Med. J. Aust., 2: 573-574.

    MULLER, E. (1937) [Chloracne (caused by chlorinated benzenes).
    Inaugural Dissertation,] Speyer am Rhein, Pilger-Druckerei (Thesis,
    Friedrich-Wilhelm University, Breslau).


    MURRAY, F.J., SMITH, F.A., NITSCHKE, K.D., HUMISTON, C.G., KOCIBA,
    R.J., & SCHWETZ, B.A. (1979) Three-generation reproduction study of
    rats given 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the diet.
    Toxicol. appl. Pharmacol., 50: 241-252.

    NAGARKATTI, P.S., SWEENEY, G.D., GAULDIE, J., & CLARK, D.A. (1984)
    Sensitivity to suppression of cytotoxic T cells generation by
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in dependent on the Ah
    genotype of the murine host. Toxicol. appl. Pharmacol., 72:
    169-176.

    NAGAYAMA, J., KURATSUNE, M., & MASUDA, Y. (1976) Determination of
    chlorinated dibenzofurans in kanechlors and "Yusho oil". Bull.
    environ. Contam. Toxicol., 15: 9-13.

    NAGAYAMA, J., NISHIZUMI, M., & MASUDA, Y. (1979) [Subacute toxicity of
    polychlorinated dibenzofurans in mice.] Fukuoka Igakuzassh, 70:
    109-113 (in Japanese).

    NAGAYAMA, J., TOKUDOME, S., & KURATSUNE, M. (1980) Transfer of
    polychlorinated dibenzofurans to the foetuses and offspring of mice.
    Food Cosmet. Toxicol., 18: 153-157.

    NAGAYAMA, J., KUROKI, H., MASUDA, Y., & KURATSUME, M. (1983) A
    comparative study of polychlorinated dibenzofurans, poly-chlorinated
    biphenyls and 2,3,7,8-tetrachlorodibenzo-p-dioxin on aryl hydrocarbon
    hydroxylase inducing potency in rats. Arch. Toxicol., 53: 177-184.

    NAGAYAMA, J., KUROKI, H., MASUDA, Y., HANDA, S., & KURATSUNE, M.
    (1985a) Genetically mediated induction of aryl hydrocarbon hydroxylase
    activity in mice by polychlorinated dibenzofuran isomers and
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch. Toxicol., 56: 226-229.

    NAGAYAMA, J., KIYOHARA, C., KURATSUNE, M., & MASUDA, Y. (1985b)
    Induction of aryl hydrocarbon hydroxylase activity in human
    lymphoblastoid cells by chlorinated dibenzofuran isomers and
    2,3,7,8-tetrachlorodibenzo-p-dioxin. In: Keith, L.H., Rappe, C., &
    Choudhary, G., ed. Chlorinated dioxins and dibenzofurans in the
    total environment II, Stoneham, Maine, Butterworth Publishers, pp.
    285-295.

    NAU, H. & BASS, R. (1981) Transfer of
    2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD) to the mouse embryo and
    fetus. Toxicology, 20: 299-308.

    NAU, H., BASS, R., & NEUBERT, D. (1986) Transfer of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) via placenta and milk, and
    postnatal toxicity in the mouse. Arch. Toxicol., 59: 36-40.

    NCI (1977) Bioassay of dibenzo-p-dioxin for possible
    carcinogenicity, Washington, DC, National Cancer Institute, 104 pp
    (Technical Report No. 122).

    NCI (1979) Bioassay of 2,7-dichlorodibenzo-p-dioxin (DCDD) for
    possible carcinogenicity, Washington, DC, National Cancer Institute,
    103 pp (Technical Report No. 123).

    NEAL, R.A., BEATTY, P.W., & GASIEWICZ, T.A. (1979) Studies of the
    mechanisms of toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).
    Ann. N.Y. Acad. Sci., 320: 204-213.

    NEBERT, D.W., ROBINSON, J.R., NIWA, A., KUMAKI, K., & POLAND, A.P.
    (1975) Genetic expression of aryl hydrocarbon hydroxylase activity in
    the mouse. J. cell. Physiol., 85: 393-414.

    NEBERT, D.W., JENSEN, N.M., PERRY, J.W., & OKA, T. (1980) Association
    between ornithine decarboxylase induction and the Ah locus in mice
    treated with polycyclic aromatic compounds. J. biol. Chem., 255:
    6836-6842.

    NESTRICK, T.J., LAMPARSKI, L.L., & STEHL, R.H. (1979) Synthesis and
    identification of the 22 tetrachlorodienzo-p-dioxin isomers by high
    performance liquid chromatography and gas chromatography. Anal.
    Chem., 51: 2273-2281.

    NESTRICK, T.J., LAMPARSKI, L.L., & TOWNSEND, D.I. (1980)
    Identification of tetrachlorodibenzo-p-dioxin isomers at the 1 ng
    level by photolytic degradation and patter recognition techniques.
    Anal. Chem., 52: 1865-1874.

    NESTRICK, T.J., LAMPARSKI, L.L., FRAWLEY, N.N., HUMMEL, R.A., KOCHER,
    C.W., MAHLE, N.H., MCCOY, J.W., MILLER, D.L., PETERS, T.L., PILLEPICH,
    J.L., SMITH, W.E., & TOBEY, S.W. (1986) Perspectives of a large scale
    environmental survey for chlorinated dioxins: Overview and soil data.
    Chemosphere, 15: 1453-1460.

    NEUBERT, D. & DILLMANN, I. (1972) Embryotoxic effects in mice treated
    with 2,4,5-trichlorophenoxyacetic acid and
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Naunyn-Schmiedeberg's Arch.
    Pharmacol., 272: 243-264.

    NIEMANN, R.A., BRUMLEY, W.C., FIRESTONE, D., & SPHON, J.A. (1983)
    Analysis of fish for 2,3,7,8-tetrachlorodibenzo-p-dioxin by electron
    capture capillary gas chromatography. Anal. Chem., 55: 1497-1504.

    NIH (1980a) Bioassay of 1,2,3,6,7,8-and
    1,2,3,7,8,9-hexa-chlorodibenzo-p-dioxin for possible
    carcinogenicity (dermal study), Bethesda, Maryland, US Department of
    Health and Human Services, National Institutes of Health (DHSS
    Publication No. (NIH) 80-1758).

    NIH (1980b) Bioassay of 1,2,3,6,7,8-and
    1,2,3,7,8,9-hexa-chlorodibenzo-p-dioxin for possible
    carcinogenicity (gavage), Bethseda, Maryland, US Department of
    Health and Human Services, National Institutes of Health (DHHS
    Publication No. (NIH) 80-1754).

    NIH (1982a) Carcinogenesis bioassay of
    2,3,7,8-tetrachloro-dibenzo-p-dioxin (CAS No 1746-01-6) in
    Osborne-Mendel rats and B6C3F1 mice (gavage study), Research
    Triangle Park, North Carolina, US National Institutes of Health,
    National Toxicology Program (NTP Technical Report Series No. 209).

    NIH (1982b) Carcinogenesis bioassay of
    2,3,7,8-tetrachloro-dibenzo-p-dioxin (CAS No 1746-01-6) in
    Swiss-Webster mice (dermal study), Research Triangle Park, North
    Carolina, US National Institutes of Health, National Toxicology
    Program (NTP Technical Report Series No. 201).

    NILSSON, C.A., ANDERSSON, K., RAPPE, C., & WESTERMARK, S.O. (1974)
    Chromatographic evidence for the formation of chloro-dioxins from
    chloro-2-phenoxyphenols. J. Chromatogr., 96: 137-147.

    NILSSON, C.A., NORSTROM, A., ANDERSSON, K., & RAPPE, C. (1978)
    Impurities in commercial products related to pentachlorophenol. In:
    Rao, K.R., ed. Pentachlorophenol: Chemistry, pharmacology and
    environmental toxicology, New York, London, Plenum Press, pp.
    313-324.

    NISBET, I.C.T. & PAXTON, M.B. (1982) Statistical aspects of
    three-generation studies of the reproductive toxicity of TCDD and
    2,4,5,-T. Am. Stat., 36(3): 290-298.

    NISHIZUMI, M. (1978) Acute toxicity of polychlorinated dibenzofurans
    in CF-1 mice. Toxicol. appl. Pharmacol., 45: 209-212.

    NISHIZUMI, M., KURATSUNE, M., & MASSUDA, Y. (1975) [Comparison of
    hyperkeratosis induced by PCBs, PCDF and PCDD application.
    Polychlorinated biphenyls, polychlorinated dibenzofuran and
    polychlorinated dibenzodioxin.] Fukuoka Asa Med., 66: 600-604 (in
    Japanese).

    NIWA, A., KUMAKI, K., & NEBERT, D.W. (1975) Induction of aryl
    hydrocarbon - hydroxylase activity in various cell cultures by
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Mol. Pharmacol., 11: 399-408.

    NOLAN, R.J., SMITH, F.A., & HEFNER, J.G. (1979) Elimination and tissue
    distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in female
    guinea pigs following a single oral dose. Tox. appl. Pharmacol.,
    48: 167 (abstract).

    NORBACK, D.H. & ALLEN, J.R. (1973) Biological responses of the
    nonhuman primate, chicken, and rat to chlorinated dibenzo-p-dioxin
    ingestion. Environ. Health Perspect., 5: 233-240.

    NORBACK, D.H., ENGBLOM, J.F., & ALLEN, J.R. (1975) Tissue distribution
    and excretion of octachlorodibenzo-p-dioxin in the rat. Toxicol.
    appl. Pharmacol., 32: 330-338.

    NORMAN, L.R., JOHNSON, E.F., & MULLER-EBERHARD, U. (1978)
    Identification of the major cytochrome P-450 form transplacentally
    induced in neonatal rabbits by 2,3,7,8-tetra-chlorodibenzo-p-dioxin.
    J. biol. Chem., 253: 8640-8647.

    NORSTROM, R.J., HALLETT, D.J., SIMON, M., & MULVIHILL, M.J. (1982)
    Analysis of Great Lakes Herring Gull eggs for
    tetra-chlorodibenzo-p-dioxins. In: Hutzinger, O., ed. Chlorinated
    dioxins and related compounds. Impact on the environment, Oxford,
    New York, Pergammon Press, pp. 173-182.

    NORSTROM, R.J., SIMON, M., & WESELOH, D.V. (1986) Long-term trends
    of PCDD and PCDF contamination in the Great Lakes. Presented at
    Dioxin 86, 6th International Symposium on Chlorinated Dioxins and
    Related Compounds, Fukuoka, Japan, 16-19 September, 1986.

    NORSTROM, R.J., SIMON, M., WHITEHEAD, P.E., KUSSAT, R., & GARRET, C.
    (1988) Level of polychlorinated dibenzo-p-dioxins (PCDDs) and
    polychlorinated dibenzofurans (PCDFs) in biota and sediments near
    potential sources of contamination in British Columbia, 1987, West
    Vancouver, British Columbia, Environment Canada (Analytical Report
    CRD-88-5).


    NRC CANADA (1981) Polychlorinated dibenzo-p-dioxins. Limitation to
    the current analytical technique, Ottawa, National Research Council
    of Canada (Publication No. NRCC/NRC 18576).

    NYGREN, M., RAPPE, C., LINDSTROM, G., HANSSON, M., BERGQVIST, P.-A.,
    MARKLUND, S., DOMELLOF, L., HARDELL, L., & OLSSON, M. (1986)
    Identification of 2,3,7,8-substituted polychlorinated dioxins and
    dibenzofurans in environmental and human samples. In: Rappe, C.,
    Choudhary, G., & Keith, L., ed. Chlorinated dioxins and
    dibenzofurans in perspective, Chelsea, Michigan, Lewis Publishers,
    pp. 17-34.

    OBERG, T. & BERGSTROM, J. (1986) Combustion test data from a Swedish
    hazardous waste incinerator. Chemosphere, 15: 2045-2048.

    OISHI, S. (1977) Influence of polychlorinated dibenzofurans (PCDFs)
    and polychlorinated biphenyls (PCBs) to serum protein components in
    rats. Bull. environ. Contam. Toxicol., 18: 773-777.

    OISHI, S. & HIRAGA, K. (1980) Effect of polychlorinated biphenyl,
    dibenzofuran and dibenzo-p-dioxin on the susceptibility of male mice
    to endotoxin. J. environ. Sci. Health, B15: 77-85.

    OISHI, S., MORITA, M., & FUKUDA, H. (1978) Comparative toxicity of
    polychlorinated biphenyls and dibenzofurans in rats. Toxicol. appl.
    Pharmacol., 43: 13-22.

    O'KEEFE, P., MESELSON, M.S., & BAUGHMAN, R. (1977) Neutral clean-up
    procedure for 2,3,7,8-tetrachlorodibenzo-p-dioxin residues in bovine
    fat and milk. J. Assoc. Off. Anal. Chem., 61: 621-626.

    O'KEEFE, P., MEYER, C., HILKER, D., ALDOVES, K., JELUS-TYROR, B.,
    DILLON, K., DONNOLLY, R., HORN, E., & SLOAN, R. (1983) Analysis of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in Great Lakes fish. Chemosphere,
    12: 325-332.

    O'KEEFE, P., SILKWORTH, J.B., BIERTHY, J.F., SMITH, R.M., DECAPRIO,
    A.P., TURNER, J.N., EADON, G., HILKER, D.R., ALDOUS, K.M., KAMINSKY,
    L.S., & COLLINS, D.N. (1985) Chemical and biological investigations of
    a transformer accident at Binghamton, NY. Environ. Health
    Perspect., 60: 201-209.

    OKEY, A.B. & VELLA, L.M. (1982) Binding of 3-methylchlo-anthrene and
    2,3,7,8-tetrachlorodibenzo-p-dioxin to a common Ah receptor site in
    mouse and rat hepatic cytosols. Eur. J. Biochem., 127: 39-47.


    OKEY, A.B., BONDY, G.P., MASON, M.E., KAHL, G.F., EISEN, H.J.,
    GUENTHNER, T.M., & NEBERT, D.W. (1979) Regulatory gene product of the
    Ah locus. J. biol. Chem., 254: 11636-11648.

    OKEY, A.B., BONDY, G.P., MASON, M.E., NEBERT, D.W., FORSTER-GIBSON,
    C.J., MUNCHAN, J., & DUFRESNE, M.J. (1980) Temperature-dependent
    cytosol-to-nucleus translocation of the Ah receptor for
    2,3,7,8-tetrachlorodibenzo-p-dioxin in continuous cell culture lines.
    J. biol. Chem., 225: 11415-11422.

    OKINO, S.T., QUATTROCHI, L.C., BARNES, H.J., OSANTO, S., GRIFFIN,
    K.J., JOHNSON, E.F., & TUKEY, R.H. (1985) Cloning and characterization
    of cDNAs encoding 2,3,7,8-tetrachlorodibenzo-p-dioxin-inducible rabbit
    mRNAs for cytochrome P-450 isozymes 4 and 6. Proc. Natl Acad. Sci.
    (USA), 82: 5310-5314.

    OLIE, K., VERMEULEN, P.L., & HUTZINGER, O. (1977)
    Chlorodibenzo-p-dioxins and chlorodibenzofurans are trace components
    of fly ash and flue gas of some municipal incinerators in the
    Netherlands. Chemosphere, 8: 455-459.

    OLIE, K., BERG, M.V.D., & HUTZINGER, O. (1983) Formation and fate of
    PCDD and PCDF from combustion processes. Chemosphere, 12: 627-636.

    OLIVER, R.M. (1975) Toxic effects of
    2,3,7,8-tetrachloro-dibenzo-1,4-dioxin in laboratory workers. Br. J.
    ind. Med., 32: 49-53.

    OLSON, J.A. & GUNNING, D. (1983) The storage form of vitamin A in rat
    liver cells. J. Nutr., 113: 2184-2191.

    OLSON, J.R. (1986) Metabolism and disposition of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in guinea pigs. Toxicol. appl.
    Pharmacol., 85: 263-273.

    OLSON, J.R., GASIEWICZ, T.A., & NEAL, R.A. (1980a) Tissue
    distribution, excretion, and metabolism of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin (TCDD) in the golden Syrian
    hamster. Toxicol. appl. Pharmacol., 56: 78-85.

    OLSON, J.R., HOLSCHER, M.A., & NEAL, R.A. (1980b) Toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in the golden syrian hamster.
    Toxicol. appl. Pharmacol., 55: 67-78.

    ONO, M., WAKIMOTO, T., TATSUKAWA, R., & MASUDA, Y. (1986)
    Polychlorinated dibenzo-p-dioxins and dibenzofurans in human adipose
    tissues of Japan. Chemosphere, 15: 1629-1634.


    ONTARIO (1985) Scientific criteria document for standard
    development. Polychlorinated dibenzo-p-dioxins (PCDDs) and
    polychlorinated dibenzofurans (PCDFs), Toronto, Ontario Ministry of
    the Environment (Report No. 4-84).

    ONTARIO (1986) Drinking water survey St. Clair river - Detroit river
    area, Toronto, Ontario Ministry of the Environment.

    ORR, J.B. & RICHARDS, M.B. (1934) Growth and vitamin A deficiency.
    Biochem. J., 28: 1259-1273.

    OSBORNE, R. & GREENLEE, W.F. (1985)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin (TCDD) enhances terminal
    differentiation of cultured human epidermal cells. Toxicol. appl.
    Pharmacol., 77: 434-443.

    OSBORNE, R., DOLD, K.M., & GREENLEE, W.F. (1984) Cell culture models
    to study mechanisms of toxicity of chlorinated aromatic compounds to
    skin and thymus. Chem. Ind. Inst. Toxicol., 4: 2-7.

    OTT, M.G., HOLDER, B.B., & OLSON, R.D. (1980) A mortality analysis of
    employees engaged in the manufacture of 2,4,5-trichlorophenoxyacetic
    acid. J. occup. Med., 22: 47-50.

    PAASIVIRTA, J., ENQVIST, J., RAISANEN, S., & PAASIVUO, P. (1977) On
    the limit of detection of TCDD in gas chromatography. Chemosphere,
    6: 355.

    PALAUSKY, J., HARWOOD, J.J., CLEVENGER, T.E., KAPILA, S., & YANDERS,
    A.F. (1986) Disposition of tetrachlorodibenzo-p-dioxin in soil. In:
    Rappe, C., Choudhary, G., & Keith, L., ed. Chlorinated dioxins and
    dibenzofurans in perspective, Chelsea, Michigan, Lewis Publishers,
    pp. 211-223.

    PATTERSON, D.G., Jr, HOFFMAN, R.E., NEEDHAM, L.L., ROBERTS, D.W.,
    BAGBY, J.R., PIRKLE, J.L., FALK, H., SAMPSON, E.J., & HOUK, V.N.
    (1986) 2,3,7,8-Tetrachlorodibenzo-p-dioxin levels in adipose tissue of
    exposed and control persons in Missouri. J. Am. Med. Assoc., 256:
    2683-2686.

    PATTERSON, J.M., MCHENRY, E.W., & CRANDALL, W.A. (1942) The
    physiological properties of vitamin A. 1. A specific effect upon body
    weight and body composition in the Albino rat. Biochem. J., 36:
    792-794.

    PAZDERNIK, T.L. & ROZMAN, K.K. (1985) Effect of thyroidectomy and
    thyroxine on 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced
    immunotoxicity. Life Sci., 36: 695-703.


    PAZDEROVA-VEJLUPKOVA, J., LUKAS, E., NEMCOVA, M., SPACILOVA, M.,
    JIRASEK, L., KALENSKY, J., JOHN, J., JIRASEK, A., & PICKOVA, J. (1974)
    [Chronic intoxication by chlorinated hydrocarbons formed during the
    production of sodium 2,4,5-trichlorophenoxyacetate.] Pracov. Lek.,
    26: 332-339 (in Czech).

    PAZDEROVA-VEJLUPKOVA, J., LUKAS, E., NEMCOVA, M., PICKOVA, J., &
    JIRASEK, L. (1980) [Chronic poisoning by
    2,3,7,8-tetra-chlorodibenzo-p-dioxin.] Pracov. Lek., 32: 204-209
    (in Czech).

    PAZDEROVA-VEJLUPKOVA, J., NEMCOVA, M., PICKOVA, J., JIRASEK, L., &
    LUKAS, E. (1981) The development and prognosis of chronic intoxication
    by tetrachlorodibenzo-p-dioxin in men. Arch. environ. Health, 36:
    5-11.

    PETERSON, R.E., MADHUKAR, B.V., YANG, K.H., & MATSUMURA, F. (1979a)
    Depression of adenosine triphosphatase activities in isolated liver
    surface membranes of 2,3,7,8-tetrahlorodibenzo-p-dioxin-treated rats:
    correlation with effects on ouabain biliary excretion and bile flow.
    J. Pharmacol. exp. Ther., 210: 275-210.

    PETERSON, R.E., HAMADA, N., YANG, K.H., & MADHUKAR, B.V. (1979b)
    Reversal of 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced depression of
    ouabain biliary excretion by pregnenolone-16-beta-carbonitrile and
    spironolactone in isolated perfused rat livers. Toxicol. appl.
    Pharmacol., 50: 407-416.

    PHILIPPI, M., SCHMID, J., WIPF, H.K., & HUTTER, R. (1982) A microbial
    metabolite of TCDD. Experientia (Basel), 38: 659-661.

    PIPER, W.N., ROSE, J.Q., & GEHRING, P.J. (1973) Excretion and tissue
    distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Adv.
    Chem. Ser., 120: 85-91.

    PITOT, H.C. & SIRICA, A.E. (1980) The stages of initiation and
    promotion in hepatocarcinogenesis. Biochim. Biophys. Acta, 605:
    191-215.

    PITOT, H.C., GOLDSWORTHY, T., CAMPBELL, H.A., & POLAND, A. (1980)
    Quantitative evaluation of the promotion by
    2,3,7,8-tetrachlorodibenzo-p-dioxin of hepatocarcinogenesis from
    diethylnitrosamine. Cancer Res., 40: 3616-3620.

    POCCHIARI, F. (1978) 2,3,7,8-Tetrachlorodibenzo-para-dioxin
    decontamination. Ecol. Bull. (Stockholm), 27: 67-70.


    POCCHIARI, F., DIDOMENICO, A., SILANO, V., & ZAPPONI, G. (1983)
    Environmental impact of the accidental release of
    tetrachlorodibenzo-p-dioxn (TCDD) at Seveso (Italy). In: Coulston, F.
    & Pocchiari, F., ed. Accidental exposure to dioxins. human health
    aspects, New York, London, Academic Press, pp. 5-35.

    POELLINGER, L. & GULLBERG, D. (1985) Characterization of the
    hydrophobic properties of the receptor for
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Mol. Pharmacol., 27:
    271-276.

    POELLINGER, L., KURL, R.N., LUND, J., GILLNER, M., CARLSTEDT-DUKE, J.,
    HOGBERG, B., & GUSTAFSSON. J.-A. (1982) High-affinity binding of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in cell nuclei from rat liver.
    Biochem. Biophys. Acta, 714: 516-523.

    POELLINGER, L., LUND, J., HANSSON, L.-A., & GUSTAFSSON, J.-A. (1983)
    Physicochemical characterization of specific and non-specific
    polyaromatic hydrocarbon binders in rat and mouse liver cytosol. J.
    biol. Chem., 258: 13535-13542.

    POHJANVIRTA, R. & TUOMISTO, J. (1986) Han/Wistar rats are
    exceptionally resistant to TCDD. II. Arch. Toxicol., 11 (Suppl.):
    344-347.

    POHLAND, A.E. & YANG, G.C. (1972) Preparation and characterization of
    chlorinated dibenzo-p-dioxins. J. agric. food Chem., 20(6):
    1093-1099.

    POIGER, H. & BUSER, H.R. (1983) Structure elucidation of mammalian
    TCDD-metabolites. In: Tucker, R.E., Young, A.L., & Gray, A.P., ed.
    Human and environmental risks of chlorinated dioxins and related
    compounds, New York, London, Plenum Press, pp. 483-492.

    POIGER, H. & SCHLATTER, Ch. (1979) Biological degradation of TCDD in
    rats. Nature (London), 281: 706-707.

    POIGER, H. & SCHLATTER, Ch. (1980) Influence of solvents and
    adsorbents on dermal and intestinal absorption of TCDD. Food Cosmet.
    Toxicol., 18: 477-481.

    POIGER, H. & SCHLATTER, Ch. (1985) Influence of phenobarbital and TCDD
    on the hepatic metabolism of TCDD in the dog. Experientia (Basel),
    41: 376-378.

    POIGER, H. & SCHLATTER, Ch. (1986) Pharmacokinetics of 2,3,7,8-TCDD in
    man. Chemosphere, 15: 1489-1494.

    POIGER, H., BUSER, H.R., WEBER, H., ZWEIFEL, U., & SCHLATTER, Ch.
    (1982) Structure elucidation of mammalian TCDD-metabolites.
    Experientia (Basel), 38: 484-486.

    POIGER, H., BUSER, H.R., & SCHLATTER, Ch. (1984) The metabolism of
    2,3,7,8-tetrachlorodibenzofuran in the rat. Chemosphere, 13:
    351-357.

    POLAND, A. & GLOVER, E. (1973a) Studies on the mechanism of toxicity
    of the chlorinated dibenzo-p-dioxins. Environ. Health Perspect.,
    5: 245-251.

    POLAND, A. & GLOVER, E. (1973b) 2,3,7,8-Tetrachlorodibenzo-p-dioxin:
    A potent inducer of aminolevulinic acid synthetase. Science, 179:
    476-477.

    POLAND, A. & GLOVER, E. (1973c) Chlorinated dibenzo-p-dioxins: Potent
    inducers of aminolevulinic acid synthetase and aryl hydrocarbon
    hydroxylase. II. A study of the structure-activity relationship.
    Mol. Pharmacol., 9: 736-747.

    POLAND, A. & GLOVER, E. (1974a) The induction of aryl hydrocarbon
    hydroxylase by 2,3,7,8-tetrachlorodibenzo-p-dioxin: Evidence for a
    receptor mutation in genetically non-responsive mice. Pharmacologist,
    16: 240 (Abstract 282).

    POLAND, A. & GLOVER, E. (1974b) Comparison of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin, a potent inducer of aryl
    hydrocarbon hydroxylase, with 3-methylcholanthrene. Mol. Pharmacol.,
    10: 349-359.

    POLAND, A. & GLOVER, E. (1974c) Genetic expression of aryl hydrocarbon
    hydroxylase activity. Induction of monooxygenase activities and
    cytochrome P1-450 formation by 2,3,7,8-tetra-chlorodibenzo-p-dioxin in
    mice genetically "nonresponsive" to other aromatic hydrocarbons. J.
    biol. Chem., 249: 5599-5606.

    POLAND, A. & GLOVER, E. (1975) Genetic expression of arylhydrocarbon
    hydroxylase by 2,3,7,8-tetrachlorodibenzo-p-dioxin: Evidence for a
    receptor mutation in genetically non-responsive mice. Mol.
    Pharmacol., 11: 389-398.

    POLAND, A. & GLOVER, E. (1979) An estimate of the maximum in vivo
    covalent binding of 2,3,7,8-tetrachlorodibenzo-p-dioxin to rat liver
    protein, ribosomal RNA, and DNA. Cancer Res., 39: 3341-3344.

    POLAND, A. & GLOVER, E. (1980) 2,3,7,8-Tetrachlorodibenzo-p-dioxin:
    Segregation of toxicity with the Ah-locus. Mol. Pharmacol., 17:
    86-94.


    POLAND, A. & KENDE, A. (1976) 2,3,7,8-Tetrachlorodibenzo-p-dioxin:
    Environmental contaminant and molecular probe. Fed. Proc., 35:
    2404-2411.

    POLAND, A. & KNUTSON, J.C. (1982) 2,3,7,8-Tetrachlorodibenzo-p-dioxin
    and related halogenated aromatic hydrocarbons: Examination of the
    mechanisms of toxicity. Annu. Rev. Pharmacol. Toxicol., 22:
    517-554.

    POLAND, A., SMITH, D., METTER, G., & POSSICK, P. (1971) A health
    survey of workers in a 2,4-D and 2,4,5-T plant. Arch. environ.
    Health, 22: 316-327.

    POLAND, A., GLOVER, E., & KENDE, A.S. (1976) Stereospecific, high
    affinity binding of 2,3,7,8-tetrachlorodibenzo-p-dioxin by hepatic
    cytosol. Evidence that the binding species is receptor for induction
    of aryl hydrocarbon hydroxylase. J. biol. Chem., 251: 4936-4946.

    POLAND, A., PALEN, D., & GLOVER, E. (1982) Tumor promotion by TCDD in
    HRS/J mice. Nature (Lond.), 300(5889): 271-273.

    POLI, A., FRANCESCHINI, G., PUGLISI, L., & SIRTORI, C.R. (1980)
    Increased total and high density lipoprotein cholesterol with
    apoprotein changes resembling streptozotocin diabetes in
    tetrachlorodibenzodioxin (TCDD) treated rats. Biochem. Pharmacol.,
    29: 835-838.

    POTTER, C.L., SIPES, I.G., & RUSSELL, D.H. (1982) Inhibition of
    ornithine decarboxylase activity by
    2,3,7,8-tetrachloro-dibenzo-p-dioxin. Biochem. Pharmacol., 31:
    3367-3371.

    POTTER, C.L., SIPES, I.G., & RUSSELL, D.H. (1983) Hypothyroxinemia and
    hypothermia in rats in response to 2,3,7,8-tetrachlorodibenzo-p-dioxin
    administration. Toxicol. appl. Pharmacol., 69: 89-95.

    POTTER, C.L., MENAHAN, L.A., & PETERSON, R.E. (1986a) Relationship of
    alterations in energy metabolism to hypophagia in rats treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Fundam. appl. Toxicol., 6:
    89-97.

    POTTER, C.L., MOORE, R.W., INHORN, S.L., HAGEN, T.C., & PETERSON, R.E.
    (1986b) Thyroid status and thermogenesis in rats treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    84: 45-55.

    PRATT, R.M., DENCKER, L., & DIEWERT, V.M. (1984)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin-induced cleft palate in the mouse:
    Evidence for alterations in palatal shelf fusion. Teratog.
    Carcinogen. Mutagen., 4: 427-436.



    PRESTON, B.D., VAN MILLER, P.J., MOORE, R.W., & ALLEN, J.R. (1981)
    Promoting effects of polychlorinated biphenyls (Aroclor 1254) and
    polychlorinated dibenzofuran-free Aroclor 1254 on
    diethylnitrosamine-induced tumorigenesis in the rat. J. Natl Cancer
    Inst., 66: 509-515.

    PUHVEL, S.M., SAKAMOTO, M., ERTL, D.C., & REISNER, R.M. (1982)
    Hairless mice as models for chloracne: A study of cutaneous changes
    induced by topical application of established chloracnegens.
    Toxicol. appl. Pharmacol., 64: 492-503.

    PUHVEL, S.M., ERTL, D.C., & LYNBERG, C.A. (1984) Increased epidermal
    transglutaminase activity following
    2,3,7,8-tetra-chlorodibenzo-p-dioxin: In vivo and in vitro studies
    with mouse skin. Toxicol. appl. Pharmacol., 73: 42-47.

    PUHVEL, S.M., SAKAMOTO, M., & REISNER, R.M. (1986) Localization of
    TCDD in hairless mouse skin. Chemosphere, 15: 2065-2067.

    QUILLEY, C.P. & RIFKIND, A.B. (1986) Prostaglandin release by the
    chick embryo heart is increased by
    2,3,7,8-tetrachloro-dibenzo-p-dioxin and by other cytochrome P-448
    inducers. Biochem. biophys. Res. Commun., 136(2): 582-589.

    RAMSEY, J.C., HEFNER, J.G., KARBOWSKI, R.J., BRAUN, W.H., & GEHRING,
    P.J. (1982) The in vivo biotransformation of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the rat. Toxicol.
    appl. Pharmacol., 65: 180-184.

    RAPPE, C. (1978a) Chemical background of the phenoxy acids and
    dioxins. Ecol. Bull. (Stockholm), 27: 28-30.

    RAPPE, C. (1978b) Decontamination of products formed during the
    industrial preparation of 2,4,5-trichlorophenol. In: Cattabeni, F.,
    Cavallero, A., & Galli, G., ed. Dioxin: toxicological and chemical
    aspects, New York, SP Medical and Scientific Books, pp. 179-183.

    RAPPE, C. (1984) Analysis of polychlorinated dioxins and furans.
    Environ. Sci. Technol., 18: 78A-90A.

    RAPPE C. (1985) Problems in analysis of PCDDs and PCDFs and presence
    of these compounds in human milk, Copenhagen, WHO Regional Office for
    Europe (ICP/CEH/501/05/7).

    RAPPE, C. (1987) Global distribution of polychlorinated dioxins
    (PCDDs) and dibenzofurans (PCDFs). In: Solving hazardous waste
    problems: Learning from dioxins, New York, American Chemical
    Society, pp. 20-23 (ACS Symposium Series No. 338).

    RAPPE, C. & BUSER, H.R. (1980) Chemical properties and analytical
    methods. In: Kimbrough, R.D., ed. Halogenated biphenyls, terphenyls,
    naphthalenes, dibenzodioxins and related products, Amsterdam,
    Oxford, New York, Elsevier Science Publishers, pp. 41-80.

    RAPPE, C. & KJELLER, L.-O. (1987) PCDDs and PCDFs in environ-mental
    samples air, particulates, sediments and soil. Chemosphere, 16:
    1775-1780.

    RAPPE, C. & NYGREN, M. (1984) Chemical analysis of human samples.
    Identification and quantification of polychlorinated dioxins and
    dibenzofurans. In: de Serres, F.J. & Pero, R.W., ed. Individual
    susceptibility to genotoxic agents in the human population, New
    York, London, Plenum Press, pp. 305-314.

    RAPPE, C., BUSER, H.R., & BOSHARDT, H.P. (1978) Identification and
    quantification of polychlorinated dibenzo-p-dioxins (PCDDs) and
    dibenzofurans (PCDFs) in 2,4,5-T-ester formulations and herbicide
    orange. Chemosphere, 7: 431-438.

    RAPPE, C., BUSER, H.R., KUROKI, H., & MASUDA, Y. (1979) Identification
    of polychlorinated dibenzofurans (PCDFs) retained in patients with
    Yusho. Chemosphere, 4: 259-266.

    RAPPE, C., BUSER, H.R., STALLING, D.L., SMITH, L.M., & DOUGHERTY, R.C.
    (1981) Identification of polychlorinated dibenzofurans in
    environmental samples. Nature (Lond.), 292: 524-526.

    RAPPE, C., MARKLUND, S., BERGQVIST, P.-A., & HANSSON, M. (1983a)
    Polychlorinated dibenzo-p-dioxins, dibenzofurans and other polynuclear
    aromatics formed during incineration and polychlorinated biphenyl
    fires. In: Choudhary, G., Keith, L.H., & Rappe, C., ed. Chlorinated
    dioxins and dibenzofurans in the total environment, Boston,
    Butterworth Publishers, pp. 99-124.

    RAPPE, C., MARKLUND, S., NYGREN, M., & GARA, A. (1983b) Parameters for
    identification and confirmation in trace analyses of polychlorinated
    dibenzo-p-dioxins and dibenzo-furans. In: Choudhary, G., Keith, L.H.,
    & Rappe, C., ed., Chlorinated dioxins and dibenzofurans in the total
    environment, Boston, Butterworth Publishers, pp. 240-259.

    RAPPE, C., NYGREN, M., BUSER, H.R., MASUDA, Y., KUROKI, H., & CHEN,
    P.H. (1983c) Identification of polychlorinated dioxins (PCDDs) and
    dibenzofurans (PCDFs) in human samples, occupational exposure and
    Yusho patients. In: Tucker, R.E., Young, A.L., & Gray, A.P., eds.
    Human and environmental risks of chlorinated dioxins and related
    compounds, New York, London, Plenum Press, pp. 241-254.

    RAPPE, C., BERGQVIST, P.-A., & MARKLUND, S. (1985a) Analysis of
    polychlorinated dibenzofurans and dioxins in ecological samples. In:
    Keith, L.H., Rappe, C., & Choudhary, G., ed. Chlorinated dioxins and
    dibenzofurans in the total environment II, Boston, Butterworth
    Publishers, pp. 135-138.

    RAPPE, C., MARKLUND, S., KJELLER, L-O., BERGQVIST, P.-A., & HANSSON,
    M. (1985b) Composition of polychlorinated dibenzofurans (PCDF) formed
    in PCB fires. In: Keith, L.H., Rappe, C., & Choudhary, G., ed.
    Chlorinated dioxins and dibenzofurans in the total environment II,
    Boston, Butterworth Publishers, p. 401-424.

    RAPPE, C., NYGREN, M., MARKLUND, S., KJELLER, L.-O., BERGQVIST, P.A.,
    & HANSON, M. (1985c) Assessment of human exposure to polychlorinated
    dibenzofurans and dioxins. Environ. Health Perspect., 60: 303-304.

    RAPPE, C., KJELLER, L.-O., & MARKLUND, S. (1985d) PCDF isomers and
    isomer levels found in PCBs. In: Komai, R.Y. & Addis, G., ed.
    Proceedings of a Workshop on PCB By-Product Formation, Palo Alto,
    California, 4-6 December, 1984, Palo Alto, California, Electric
    Power Research Institute, pp. 20-23.

    RAPPE, C., KJELLER, L.-O., MARKLUND, S., & NYGREN, M. (1986a)
    Electrical PCB accident, an update. Chemosphere, 15: 1291-1295.

    RAPPE, C., NYGREN, M., HANSSON, M., & KAHN, P.C. (1986b) Analysis of
    adipose tissue and blood samples from Vietnam veterans. Clean-up,
    analysis and quality control, P 41, In: Dioxin 86, 6th International
    Symposium on Chlorinated Dioxins and Related Compounds, Fukuoka,
    Japan, 16-19 September, 1986.

    RAPPE, C., ANDERSSON, R., BERGQVIST, P.-A., BROHEDE, C., HANSSON, M.,
    KJELLER, L.-O., LINDSTROM, G., MARKLUND, S., NYGREN, M., SWANSON,
    S.E., TYSKLIND, M., & WIBERG, K. (1987) Overview on environmental fate
    of chlorinated dioxins and dibenzofurans, sources, levels and isomeric
    pattern in various matrices. Chemosphere, 16: 1603-1618.

    RAPPE, C., NYGREN, M., LINDSTROM, G., BUSER, H.R., BLASER, O., &
    WUTHRICH, C. (1987) Polychlorinated dibenzofurans and
    dibenzo-p-dioxins and other chlorinated contaminants in cow milk from
    various locations in Switzerland. Environ. Sci, Technol., 21:
    964-970.

    REGGIANI, G. (1978) Medical problems raised by the TCDD contamination
    in Seveso, Italy. Arch. Toxicol., 40: 161-188.


    REGGIANI, G. (1980a) Acute human exposure to TCDD in Seveso, Italy.
    J. Toxicol. environ. Health, 6: 27-43.

    REGGIANI, G. (1980b) Toxicology of TCDD and related compounds:
    observation in man. In: Hutzinger, O., Frei, R.W., Merian, E., &
    Pocchiari, F., ed. Chlorinated dioxins and related compounds. Impact
    on the environment, Oxford, New York, Pergamon Press, pp. 463-493.

    REGGIANI, G. (1983a) Anatomy of TCDD spill: The Seveso accident.
    Curr. Dev., 2: 269-341.

    REGGIANI, G. (1983b) An overview on the health effects of halogenated
    dioxins and related compounds - the Yusho and Taiwan episodes. In:
    Coulston, F. & Pocchiari, F., ed. Accidental exposure to dioxins.
    Human health aspects, New York, London, Academic Press, pp. 39-67.

    REGIONE LOMBARDIA (1984) Final report and recommendations of the 6th
    International Steering Committee Meeting, Milan, 19-21 February,
    1984, Milan, Regione Lombardia, pp. 1-17.

    RICE, R.H. & CLINE, P.R. (1984) Opposing effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin and hydrocortisone on growth and
    differentiation of cultured malignant human keratinocytes.
    Carcinogenesis, 5(3): 367-371.

    RICE, R.H., CLINE, P.R., & COE, E.L. (1983) Mutually antagonistic
    effects of hydrocortisone and retinylacetate on envelope competence in
    cultured malignant human keratinocytes. J. invest. Dermatol., 81:
    176s-178s.

    RICHERT, J., VON (1962) [On neurological complications in a case of
    chlorinated hydrocarbon poisoning accompanied by chloracne.]
    Nervenarzt, 33(4): 180-184 (in German).

    RIKANS, L.E., GIBSON, D.D., & MCCAY, P.B. (1979) Evidence for the
    presence of cytochrome P-450 in rat mammary gland. Biochem.
    Pharmacol., 28: 3039-3042.

    RISSE-SUNDERMAN, A. (1959) [Intoxication by chloro-aromatics,]
    Cologne-Lindenburg, University of Cologne, Department of Dermatology
    (Thesis) (in German).

    RIZZARDINI, M., ROMANO, M., TURSI, F., SALMONA, F., VECCHI, A.,
    SIRONI, M., GIZZI, F., BENFENATI, E., GARATTINI, S., & FANELLI, R.
    (1983) Toxicological evaluation of urban waste incinerator emissions.
    Chemosphere, 12: 559-564.


    ROBERTSON, L.W., REGEL, U., FILSER, J.G., & OESCH, F. (1985) Absence
    of lipid peroxidation as determined by ethane exha-lation in rats
    treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Arch.
    Toxicol., 57: 13-16.

    ROGERS, A.M., ANDERSEN, M.E., & BACK, K.C. (1982) Mutagenicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin and perfluoro-n-decanoic acid in
    L5178Y mouse-lymphoma cells. Mutat. Res., 105: 445-449.

    ROMKES, K., PISKORSKA-PLISZCZYNSKA, J., & SAFE, S. (1987) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on hepatic and uterine estrogen
    receptor levels in rats. Toxicol. appl. Pharmacol., 87: 306-314.

    ROSE, J.Q., RAMSEY, J.C., WENTZLER, T.H., HUMMEL, R.A., & GEHRING,
    P.J. (1976) The fate of 2,3,7,8-tetrachlorodibenzo-p-dioxin following
    single and repeated oral doses to the rat. Toxicol. appl.
    Pharmacol., 36: 209-226.

    ROZMAN, K. (1984) Hexadecane increases the toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD): Is brown adipose tissue
    the primary target in TCDD-induced wasting syndrome? Biochem. biophys.
    Res. Commun., 125(3): 996-1004.

    ROZMAN, K., ROZMAN, T., & GREIM, H. (1984) Effect of thyroidectomy and
    thyroxine on 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) induced
    toxicity. Toxicol. appl. Pharmacol., 72: 372-376.

    ROZMAN, K., ROZMAN, T., SCHEUFLER, E., PAZDERNIK, T., & GREIM, H.
    (1985a) Thyroid hormones modulate the toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J. Toxicol. environ.
    Health, 16: 481-491.

    ROZMAN, K., HAZELTON, G.A., KLAASSEN, C.D., ARLOTTO, M.P., &
    PARKINSON, A. (1985b) Effect of thyroid hormones on liver microsomal
    enzyme induction in rats exposed to
    2,3,7,8-tetra-chlorodibenzo-p-dioxin. Toxicology, 37: 51-63.

    ROZMAN, K., PEREIRA, D., & IATROPOULOS, M.J. (1986) Histo-pathology of
    interscapular brown adipose tissue, thyroid, and pancreas in
    2,3,7,8-tetra-chlorodibenzo-p-dioxin (TCDD)-treated rats. Toxicol.
    appl. Pharmacol., 82: 551-559.

    RYAN, J.J., LAU, P.-Y., PILON, J.C., & LEWIS, D. (1983a)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin and
    2,3,7,8-tetrachloro-dibenzofuran residues in great lakes commercial
    and sport fish. In: Choudhary, G., Keith, L.H., & Rappe, C., ed.
    Chlorinated dioxins and dibenzofurans in the total environment,
    Boston, Buttersworth Publishers, pp. 87-97.


    RYAN, J.J., PILON, J.C., CONACHER, H.B.S., & FIRESTONE, D. (1983b)
    Inter-laboratory study on determination of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in fish. J. Assoc. Off. Anal.
    Chem., 66: 700-707.

    RYAN, J.J., LIZOTTE, R., SAKUMA, T., & MORI, B. (1985) Chlorinated
    dibenzo-p-dioxins, chlorinated dibenzofurans and pentachlorophenol in
    Canadian chicken and pork samples. J. agric. food Chem., 33:
    1021-1026.

    RYAN, J.J., SCHECTER, A., SUN, W.-F., & LIZOTTE, R. (1986)
    Distribution of chlorinated dibenzo-p-dioxins and chlorinated
    dibenzofurans in human tissues from the general population. In: Rappe,
    C., Choudhary, G., & Keith, L., ed. Chlorinated dioxins and
    dibenzofurans in perspective, Chelsea, Michigan, Lewis Publishers,
    pp. 3-16.

    RYHAGE, R. (1964) Use of mass spectrometer as a detector and analyzer
    for effluents emerging from high temperature gas liquid chromatography
    columns. Anal. Chem., 36: 759-764.

    SACCHI, G.A., VIGANO, P., FORTUNATI, G., & COCUCCI, S.M. (1986)
    Accumulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin from soil and
    nutriculture solution by bean and maize plants. Experientia (Basel),
    42: 586-588.

    SAFE, S.H. & SAFE, L.M (1984) Synthesis and characterization of
    twenty-two purified polychlorinated dibenzofuran congeners. J.
    agric. food Chem., 32: 68-71.

    SAFE, S.H., MASON, G., KEYS, B., FARRELL, K., ZMUDZKA, B., SAWYER, T.,
    PISKORSKA-PLISZEZYNSKA, J., SAFE, L.M., ROMKES, M., & BANDIERA, S.
    (1986) Polychlorinated dibenzo-p-dioxins and dibenzofurans:
    correlation between in vitro and in vivo structure-activity
    relationships (SARs). Chemosphere, 15: 1725-173.

    SAFE, S.H., MASON, G., FARRELL, K., KEYS, B., PISKORSKA-PLISZCZYNSKA,
    J., MADGE, J.A., & CHITTIM, B. (1987) Validation of in vitro bioassays
    for 2,3,7,8-TCDD equivalents. Chemosphere, 16: 1723-1728.

    SAI (SYSTEMS APPLICATIONS, INC.) (1980) Human exposure to
    atmospheric concentration of selected chemicals, Springfield,
    Virginia, National Technical Information Service, Vol. 1 (Report
    prepared for US Environmental Protection Agency, Research Triangle
    Park, North Carolina) (PB 81-193252).

    SANDERMANN, W. (1984a) [Dioxin. The history of the discovery of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD, dioxin, Seveso poison.]
    Naturwiss. Rundsch., 37(5): 173-178 (in German).

    SANDERMANN, W. (1984b) [Breakdown of dioxin by beta-radiation. The
    history of the discovery of 2,3,7,8-tetrachlorodibenzo-p-dioxin
    (Second communication).] Naturwiss. Rundsch., 37(11): 445-446 (in
    German).

    SANDERMANN, W., STOCKMANN, H., & GASTEN, R. (1957) [On the pyrolysis
    of pentachlorphenol.] Chem. Ber., 90: 690-692 (in German).

    SANGER, V.L., SCOTT, L., HAMDY, A., GALE, C., & POUNDEN, W.D. (1958)
    Alimentary toxemia in chickens. J. Am. Vet. Med. Assoc., 133:
    172-176.

    SARNA, L.P., HODGE, P.E., & WEBSTER, G.R.B. (1984) Octanol-water
    partition coefficients of chlorinated dioxins and dibenzofurans by
    reversed-phase HPLC using several C18 columns. Chemosphere, 13:
    975-983.

    SAWAHATA, T., OLSON, J.R., & NEAL, R.A. (1982) Identification of
    metabolites of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) formed on
    incubation with isolated rat hepatocytes. Biochem. biophys. Res.
    Commun., 105: 341-346.

    SAWYER, T.W. & SAFE, S. (1985) In vitro AHH induction by
    polychlorinated biphenyl and dibenzofuran mixtures: Additive effects.
    Chemosphere, 14: 79-84.

    SAWYER, T.W., JONES, D., ROSANOFF, K., MASON, G.,
    PISKORSKA-PLISZCZYNSKA, J., & SAFE, S. (1986) The biologic and toxic
    effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin in chickens.
    Toxicology, 39: 197-206.

    SCHANTZ, S.L., BARSOTTI, D.A., & ALLEN, J.R. (1978) Toxicological
    effects produced in nonhuman primates chronically exposed to fifty
    parts per trillion 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).
    Toxicol. appl. Pharmacol., 48(1): A180.

    SCHECTER, A., WEERASINGHE, W.C.A., GROSS, M., & DEKIN, A. (1986a)
    Human tissue levels of PCDDs and PCDFs in over one hundred year old
    frozen Eskimo tissue, meat, and California redwood charcoal and their
    relation to the trace chemistry theory of dioxin origin. In: Dioxin
    86. 6th International Symposium on Chlorinated Dioxins and Related
    Compounds, Fukuoka, Japan, 16-19 September, 1986, p.135.

    SCHECTER, A.J., RYAN, J.J., & CONSTABLE, J.D. (1986b) Chlorinated
    dibenzo-p-dioxin and dibenzofuran levels in human adipose tissue and
    milk samples from the north and south of Vietnam. Chemosphere, 15:
    1613-1620.

    SCHILLER, C.M., WALDEN, R., & SHOAF, C.R. (1982) Studies on the
    mechanism of 2,3,7,8-tetrachlorodibenzo-p-dioxin toxicity: Nutrient
    assimilation. Fed. Proc., 41: 1426 (Abstract).

    SCHILLER, C.M., ADCOCK, C.M., MOORE, R.A., & WALDEN, R. (1985) Effect
    of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and fasting on body
    weight and lipid parameters in rats. Toxicol. appl. Pharmacol.,
    81: 356-361.

    SCHILLER, C.M., ADCOCK, C.M., SGIAF, C.R., & WALDEN, R. (1986) Effects
    of adenine and its isomer 4-aminopyrazolo-(3,4-d)-pyrimidine on
    2,3,7,8-tetrachlorodibenzo-p-dioxin-induced mortality in rats.
    Toxicol. appl. Pharmacol., 84: 369-378.

    SCHOENY, R. (1982) Mutagenicity testing of chlorinated biphenyls and
    chlorinated dibenzofurans. Mutat. Res., 101: 45-56.

    SCHULZ, K.H. (1968) [On the symptoms and etiology of chloracne.]
    Arbeitsmed. Sozialmed. Arbeitshyg., 3: 25-29 (in German).

    SCHWETZ, B.A., NORRIS, J.M., SPARSCHU, G.L., ROWE, V.K., GEHRING,
    P.J., EMERSON, J.L., & GEHRING, C.G. (1973) Toxicology of chlorinated
    dibenzo-p-dioxins. Environ. Health Perspect., 5: 87-99.

    SEEFELD, M.D. & PETERSON, R.E. (1983)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin-induced weight loss: A proposed
    mechanism. In: Tucker, R.E., Young, A.L., & Gray, A.P., ed. Human
    and environmental risks of chlorinated dioxins and related
    compounds, New York, London, Plenum Press, pp. 405-412
    (Environmental Science Research Series).

    SEEFELD, M.D. & PETERSON, R.E. (1984) Digestible energy and efficiency
    of feed utilization in rats treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol. 74:
    214-222.

    SEEFELD, M.D., CORBETT, S.W., KEESEY, R.E., & PETERSON, R.E. (1984a)
    Characterization of the wasting syndrome in rats treated with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    73: 311-322.

    SEEFELD, M.D., KEESEY, R.E., & PETERSON, R.E. (1984b) Body weight
    regulation in rats treated with 2,3,7,8-tetrachloro-dibenzo-p-dioxin.
    Toxicol. appl. Pharmacol., 76: 526-536.

    SEILER, J.P. (1973) A survey on the mutagenicity of various
    pesticides. Experientia (Basel), 29: 622-623.

    SHADOFF, L.A., HUMMEL, R.A., & LAMPARSKI, L. (1977) A research for
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in an environment exposed
    annually to 2,4,5-trichlorophenoxyacetic acid ester (2,3,5-T)
    herbicides. Bull. environ. Contam. Toxicol., 18: 478-485.

    SHARMA, R.P. & GEHRING, P.J. (1979) Effects of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin (TCDD) on splenic lymphocyte
    transform-ation in mice after single and repeated exposures. Ann.
    N.Y. Acad. Sci., 320: 487-497.

    SHARMA, R.P., KOCIBA, R.J., & GEHRING, P.J. (1984) Immuno-toxicologic
    effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in rabbits. J.
    environ. Pathol. Toxicol. Oncogen., 5(4/5): 321-328.

    SHEFFIELD, A. (1985) Sources and releases of PCDDs and PCDFs to the
    Canadian environment. Chemosphere, 14: 811-814.

    SHIGEMATSU, N., ISHIMARU, S., SAITO, R., IDEDA, T., MATSUBA, K.,
    SUGIYAMA, K., & MASUDA, Y. (1978) Respiratory involvement in
    polychlorinated biphenyls poisoning. Environ. Res., 16: 92-100.

    SHIREMAN, R.B. & WEI, CHENG-I. (1986) Uptake of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin from plasma lipoproteins by
    cultured human fibroblasts. Chem.-biol. Interact., 58: 1-12.

    SHIVERICK, K.T. & MUTHER, T.F. (1983)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin (TCDD) effects on hepatic
    microsomal steroid metabolism and serum estradiol of pregnant rats.
    Biochem. Pharmacol., 32: 991-995.

    SHOAF, C.R. & SCHILLER, C.M. (1981) Studies on the mechanism of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) toxicity-lipid
    assimilation. II. Pharmacologist, 23: 176 (Abstract).

    SILKWORTH, J., MCMARTIN, D., DECAPRIO, A., REJ, R., O'KEEFE, P., &
    KAMINSKY, L. (1982) Acute toxicity in guinea pigs and rabbits of soot
    from a polychlorinated biphenyl-containing transformer fire.
    Toxicol. appl. Pharmacol., 65: 425-439.

    SIMPSON, C.F., PRITCHARD, W.R., & HARMS, R.H. (1959) An endotheliosis
    in chickens and turkeys caused by an unidentified dietary factor. J.
    Am. Vet. Med. Assoc., 134: 410-416.

    SLONECKER, P.J., PYLE, J.R., & CANTRELL, J.S. (1983) Identifi-cation
    of polychlorinated dibenzo-p-dioxin isomers by powder X-ray
    diffraction with electron capture gas chromatography. Anal. Chem.,
    55: 1543-1547.

    SMITH, A.G., FRANCIS, J.E., KAY, S.J.E., & GREIG, J.B. (1981) Hepatic
    toxicity and uroporphyrinogen decarboxylase activity following a
    single dose of 2,3,7,8-tetrachlorodibenzo-p-dioxin to mice. Biochem.
    Pharmacol., 30: 2825-2830.

    SMITH, A.G., FRANCIS, J.E., & GREIG, J.B. (1985) Continued depression
    of hepatic uroporphyrinogen decarboxylase activity caused by
    hexachlorobenzene 2,3,7,8-tetrachlorodibenzo-p-dioxin despite
    regeneration after partial hepatectomy. Biochem. Pharmacol., 34:
    1817-1820.

    SMITH, A.H., FISHER, D.O., GILES, H.J., & PEARCE, N. (1983) The New
    Zealand soft tissue sarcoma case-control study: Interview findings
    concerning phenoxyacetic acid exposure. Chemosphere, 12: 565-571.

    SMITH, F.A., SCHWETZ, B.A., & NITSCHKE, K.D. (1976) Terato-genicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin in CF-1 mice. Toxicol. appl.
    Pharmacol., 38: 517-523.

    SMITH, R.M., O'KEEFE, P.W., ALDOUS, K.M., HILKER, D.R., & O'BRIEN,
    J.E. (1983) 2,3,7,8-Tetrachlorodibenzo-p-dioxin in sediment samples
    from love canal storm sewers and creeks. Environ. Sci. Technol., 17:
    6-10.

    SODERKVIST, P., POELLINGER, L., & GUSTAFSSON, J.-A. (1986)
    Carcinogen-binding proteins in the rat ventral prostate: Specific and
    nonspecific high-affinity binging sites for benso(a)pyrene,
    3-methylcholanthrene, and 2,3,7,8-tetrachloro-dibenzo-p-dioxin.
    Cancer Res., 46: 651-657.

    SOUTHERLAND, J.H., KUYKENDAL, W.B., LAMASON, W.H. II, & OBERACKER,
    D.A. (1987) Assessment of combustion sources as emitters of
    chlorinated dioxin compounds: A report on the result of tier 4 of the
    national dioxin strategy. Chemosphere, 16: 2161-2168.

    SPARSCHU, G.L., DUNN, F.L., & ROWE, V.K. (1971) Study of the
    teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Food
    Cosmet. Toxicol., 9: 405-412.

    SPITSBERGEN, J.M., SCHAT, K.A., KLEEMAN, J.M., & PETERSON, R.E. (1986)
    Interactions of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) with immune
    responses of rainbow trout. Vet. Immunol. Immunopathol., 12:
    263-280.

    STALLING, D.L., SMITH, L.M., PETTY, J.D., HOGAN, J.W., JOHNSON, J.L.,
    RAPPE, C., & BUSER, H.R. (1983) Residues of polychlorinated
    dibenzo-p-dioxins and dibenzofurans in laurentian great lakes fish.
    In: Tucker, R.E., Young, A.L., & Gray, A.P., ed. Human and
    environmental riks of chlorinated dioxins and related compounds, New
    York, London, Plenum Press, pp. 221-240.

    STEHL, R.H. & LAMPARSKI, L.L. (1977) Combustion of several
    2,4,5-trichloro-phenoxy compounds: formation of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Science, 197: 1008-1009.

    STEHL, R.H., PAPENFUSS, R.R., BREDEWEG, R.A., & ROBERTS, R.W. (1973)
    The stability of pentachlorophenol and chlorinated dioxins to
    sunlight, heat, and combustion. Adv. Chem. Ser., 120: 119-125.

    STEHR, P.A., STEIN, G., FALK, H., SAMPSON, E., SMITH, S.M., STEINBERG,
    K., WEBB, K., AYRES, S., SCHRAMM, W., DONNELL, H.D., & GEDNEY, W.B.
    (1986) A pilot epidemiologic study of possible health effects
    associated with 2,3,7,8-tetrachloro-dibenzo-p-dioxin contaminations in
    Missouri. Arch. environ. Health, 41: 16-21.

    STEINBERG, K.K., MACNEIL, M.L., KARON, J.M., STEHR, P.A., NEESE, J.W.,
    & NEEDHAM, L.L. (1985) Assessment of
    2,3,7,8-tetrachlorodibenzo-p-dioxin exposure using a modified
    d-glucaric acid assay. J. Toxicol. environ. Health, 16: 743-752.

    STEINER, M., BORGES, H., FREEDMAN, L., & GRAY, S.J. (1968) Effects of
    starvation on the tissue composition of the small intestine in the
    rat. Am. J. Physiol., 215: 75-77.

    STEWARD, A.R. & BYARD, J.L. (1981) Induction of benzo(a)pyrene
    metabolism by 2,3,7,8-tetrachlorodibenzo-p-dioxin in primary cultures
    of adult rat hepatocytes. Toxicol. appl. Pharmacol., 59: 603-616.

    STOHS, S.J., HASSAN, M.Q., & MURRAY, W.J. (1983) Lipid peroxidation as
    a possible cause of TCDD toxicity. Biochem. biophys. Res. Commun.,
    111: 854-859.

    SUN, T.-T., SHIH, C., & GREEN, H. (1979) Keratin cytoskeletons in
    epithelial cells of internal organs. Proc. Natl Acad. Sci. (USA),
    76: 2813-2817.

    SUN, T.-T., EICHNER, R., NELSON, W.G., TSENG, S.C.G., WEISS, R.A.,
    JARVINEN, M., & WOODCOCK-MITCHELL, J. (1983a) Keratin classes:
    Molecular markers for different types of epithelial differentiation.
    J. invest. Dermatol., 81: 109-115.

    SUN, T.-T., EICHNER, R., NELSON, W.G., VIDRICH, A., &
    WOODCOCK-MITCHELL, J. (1983b) Keratin expression during normal
    epidermal differentiation. In: Seji, M. & Bernstein, I.A., ed.
    Current problems in dermatology. Normal and abnormal epidermal
    differentiation, New York, Karger, pp. 277-291.

    SUNDSTROM, G., JENSEN, S., JANSSON, B., & ERNE, K. (1979) Chlorinated
    phenoxyacetic acid derivatives and tetrachloro-dibenzo-p-dioxin in
    foliage after application of 2,4,5-trichlorophenoxyacetic acid esters.
    Arch. environ. Contam. Toxicol., 8: 441-448.

    SUSKIND, R.R. & HERTZBERG, V.S. (1984) Human health effects of 2,4,5-T
    and its toxic contaminants. J. Am. Med. Assoc., 251: 2372-2380.

    SUSKIND, R.R., CLEVELAND, F., KEENAN, C., AKIN, R., DAVIS, A., &
    KEHOE, R.A. (1953) Report on a clinical and environmental survey,
    Nitro, West Virginia, Monsanto Chemical Co..

    SUTER-HOFMANN, M. & SCHLATTER, Ch. (1985) Toxicity of particulate
    emissions from a municipal incinerator: critique of the concept of
    TCDD-equivalents. Chemosphere, 15: 1733-1743.

    SWEENEY, G.D. & JONES, K.G. (1983) Studies of the mechanism of action
    of hepatotoxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and
    related compounds. In: Tucker, R.E., Young, A.L., & Gray, A.P., ed.
    Human and environmental risks of chlorinated dioxins and related
    compounds, New York, London, Plenum Press, pp. 415-422.

    SWEENEY, G.D., JONES, K.G., COLE, F.M., BASFORD, D., & KRESTYNSKI, F.
    (1979) Iron deficiency prevents liver toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Science, 204: 332-335.

    SWIFT, L.L., GASIEWICZ, T.A., DUNN, G.D., SOULE, P.D., & NEAL, R.A.
    (1981) Characterization of the hyperlipidemia in guinea pigs induced
    by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. appl. Pharmacol.,
    59: 489-499.

    SWITZERLAND (1982) Environmental pollution due to dioxins and furans
    from chemical rubbish incineration plants, Bern, Ministry of
    Environment, Federal Swiss Government.

    TAYLOR, M.L., TIERNAN, T.O., RAMALINGAM, B., WAGEL, D.J., GARRETT,
    J.H., SOLCH, J.G., & FERGUSON, G.L. (1985) Synthesis, isolation, and
    characterization of the tetrachlorinated dibenzo-p-dioxins and other
    related compound. In: Keith, L., Rappe, C., & Choudhary, G., ed.
    Chlorinated dioxins and dibenzofurans in the total environment. II,
    Boston, Butterworth Publishers, pp. 17-35.

    TELEGINA, K.A. & BIKBULATOVA, L.I. (1970) [Affection of the follicular
    apparatus of the skin in workers occupied in production of butyl ether
    of 2,3,4-trichlorphenoxyacetic acid.] Vestn. Dermat. Venerol
    (Moscow), 44: 35-39 (in Russian).

    TELEKY, L. (1913) [Chlorine and hydrochloric acid. German Empire.]
    Wiener Arb. Geb. Soz. Med., 4: 55-56 (in German).

    TENCHINI, M.L., CRIMAUDO, C., PACCHETTI, G., MOTTURA, A., AGOSTI, S.,
    & DE CARLI, L. (1983) A comparative cytogenetic study on cases of
    induced abortions in TCDD-exposed and non-exposed women. Environ.
    Mutagen., 5: 73-85.

    THIBODEAUX, L.J. (1983) Offsite transport of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin from a production disposal
    facility. In: Choudhary, G., Keith, L.H., & Rappe, C., ed.
    Chlorinated dioxins and dibenzofurans in the total environment,
    Boston, Butterworth Publishers, pp. 75-85.

    THIESS, A.M., FRENTZEL-BEYME, R., & LINK, R. (1982) Mortality study of
    persons exposed to dioxin in a trichlorophenol-pro-cess accident that
    occurred in the BASF AG on November 17, 1953. Am. J. ind. Med., 3:
    179-189.

    THIGPEN, J.E., FAITH, R.E., MCCONNELL, E.E., & MOORE, J.A. (1975)
    Increases susceptibility to bacterial infection as a sequela of
    exposure to an environmental contaminant
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Infect. Immun., 12:
    1319-1324.

    THOMAS, P.T., & HINSDILL, R.D. (1979) The effect of perinatal exposure
    to tetrachlorodibenzo-p-dioxin on the immune response of young mice.
    Drug chem. Toxicol., 2: 77-98

    THUNBERG, T. & HAKANSSON, H. (1983) Vitamin A (retinol) status in the
    Gunn rat: the effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch.
    Toxicol., 53: 225-233.

    THUNBERG, T., AHLBORG, U.G., & JOHNSSON, H. (1979) Vitamin A (retinol)
    status in the rat after a single oral dose of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch. Toxicol., 42: 265-274.

    THUNBERG, T., AHLBORG, U.G., HAKANSSON, H., KRANTZ, C., & MONIER, M.
    (1980) Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on the hepatic
    storage of retinol in rats with different dietary supplies of vitamin
    A (retinol). Arch. Toxicol., 45: 273-285.

    THUNBERG, T., AHLBORG, U.G., & WAHLSTROM, B. (1984) Comparison between
    the effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin and six other
    compounds on the vitamin A storage, the UDP-glucuronosyltransferase
    and the aryl hydrocarbon hydroxylase activity in the rat liver.
    Arch. Toxicol., 55: 16-19.

    TIERNAN, T.O. (1983) Analytical chemistry of polychlorinated
    dibenzo-p-dioxins and dibenzofurans: A review of the current status.
    In: Choudhary, G., Keith, L., & Rappe, C., ed. Chlorinated dioxins
    and dibenzofurans in the total environment, Boston, Butterworth
    Publishers, pp. 211-237.

    TOFILON, P.J. (1980) Depressed guinea pig testicular microsomal
    cytochrome P-450 content by 2,3,7,8-tetrachloro-dibenzo-p-dioxin.
    Life Sci., 27: 871-876.

    TOFILON, P.J. & PIPER, W.N. (1982)
    2,3,7,8-Tetrachlorodibenzo-p-dioxin-mediated depression of rat
    testicular heme synthesis and microsomal cytochrome P-450. Biochem.
    Pharmacol., 31: 3663-3666.

    TOGNONI, G. & BONACCORSI, A. (1982) Epidemiological problems with
    TCDD. Drug Metab. Rev., 13: 447-469.

    TOSINE, H., SMILLIE, D., & REES, G.A.V. (1983) Comparative monitoring
    and analytical methodology for 2,3,7,8-TCDD in fish. In: Tucker, R.E.,
    Young, A.L., & Gray, A.P., ed. Human and environmental risks of
    chlorinated dioxins and related compounds, New York, London, Plenum
    Press, pp. 127-142.

    TOTH, K., SOMFAI-RELLE, S., SUGAR, J., & BENCE, J. (1979)
    Carcinogenicity testing of herbicide 2,4,5-trichlorophenoxy-ethanol
    containing dioxin and of pure dioxin in Swiss mice. Nature (Lond.),
    278: 548-549.

    TSUKAMOTO, H., MAKISUMI, S., HIROSE, H., KOJIMA, T., FUKUMOTO, H.,
    FUKOMOTO, K., KURATSUNE, M., NISHIZUMI, M., SHIBATA, M., NAGAI, J.,
    YAE, Y., SAWADA, K., FURUKAWA, M., YOSHIMURA, H., TATSUMI, K., OGURI,
    K., SHIMENO, H., KENO, K., KOBAYASHI, H., YANO, T., ITO, A., OKADA,
    T., INAGAMI, K., KOGA, T., TOMITA, Y., KOGA, T., YAMADA, Y.,
    MIYAGUCHI, M., SUGANO, M., HORI, K., TAKESHITA, K., MANAKO, K.,
    NAKAMURA, Y., & SHIGEMORI, N. (1969) [The chemical studies on detecton
    of toxic compounds in rice bran oils used by the patients of Yusho.]
    Fukuoka Acta med., 6: 496-512 (in Japanese).

    TSYRLOV, I.B., CHASOVNIKOVA, O.B., GRISHANOVA, A.YU., & LYAKHOVICH,
    V.V. (1986) Reappraisal of the liver benzpyrene hydroxylase
    synthesized de novo after treatment of rats with
    2,3,7,8-tetrachlorodibenzo-p-dioxin and 3-methylcholanthrene. FEBS
    Lett., 198: 225-228.

    TUCKER, A.N., VORE, S.J., & LUSTER, M.I. (1986) Supppression of B cell
    differentiation by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Mol.
    Pharmacol., 29: 372-377.

    TUKEY, R.H., HANNAH, R.R., NEGISHI, M., NEBERT, D.W., & EISEN, H.J.
    (1982a) The Ah locus: Correlation of intranuclear appearance of
    inducer-receptor complex with induction of cytochrome P1-450 mRNA.
    Cell, 31: 275-284.

    TUKEY, R.H., NEGISHI, M., & NEBERT, D.W. (1982b) Quantitation of
    hepatic cytochrome P1-450 mRNA with the use of a cloned DNA probe.
    Effects of various P-450 inducers in C57BL/6N and DBA/2N mice. Mol.
    Pharmacol., 22: 779-786.

    TULP, M. & HUTZINGER, O. (1978) Rat metabolism of polychlori-nated
    dibenzo-p-dioxins. Chemosphere, 9: 761-768.

    TUNG, T.T. (1973) Le cancer primaire du foie au Vietnam. Chirurgie,
    99: 427-436.

    TURNER, J.N. & COLLINS, D.N. (1983) Liver morphology in guinea pigs
    administered either pyrolysis products of a polychlorinated biphenyl
    trans former fluid or 2,3,7,8-tetrachloro-dibenzo-p-dioxin. Toxicol.
    appl. Pharmacol., 67: 417-429.

    TUTEJA, N., GONZALEZ, F.J., & NEBERT, D.W. (1985) Developmental and
    tissue-specific differential regulation of the mouse dioxin-inducible
    P1-450 and P3-450 genes. Dev. Biol., 112: 177-184.

    UMBREIT, T.H., PATEL, D., & GALLO, M.A. (1985) Acute toxicity of TCDD
    contaminated soil from an industrial site. Chemosphere, 14(6/7):
    945-947.

    UMBREIT, T.H., HESSE, E.J., & GALLO, M.A. (1986a) Bioavailability of
    dioxin in soil from a 2,4,5-T manufacturing site. Science, 232:
    497-499.

    UMBREIT, T.H., HESSE, E.J., & GALLO, M.A. (1986b) Comparative toxicity
    of TCDD contaminated soil from Times Beach, Missouri, and Newark, New
    Jersey. Chemosphere, 15: 2121-2124.


    UOTILA, P., PARKKI, M.G., & AITO, A. (1978) Quantitative and
    qualitative changes in the metabolism of benzo(a)pyrene in rat tissues
    after intragastric administration of TCDD. Toxicol. appl.
    Pharmacol., 46: 671-683.

    US EPA (1982) Environmental monitoring at Love Canal, Washington,
    DC, US Environmental Protection Agency, Office of Research and
    Development, Vol. I (EPA/600/4-82-030a).

    US EPA (1985) Health assessment document for polychlorinated
    dibenzo-p-dioxins, Washington, DC, US Environmental Protection
    Agency, Office of Health and Environmental Assessment
    (EPA/600/8-84/0146).

    US EPA (1987) Interim procedures for estimating risk associated with
    exposures to mixtures of chlorinated dibenzo-p-dioxins and
    -dibenzofurans (CDDs and CDFs). Risk assessment forum, Washington,
    DC, US Environmental Protection Agency, 48 pp (EPA/625/3-87/012).

    VAHRENHOLT, F. (1977) [Seveso - An unparalleled catastrophe.]
    Umwelt, 1: 59, 60, 62, 64 (in German).

    VAN, D.D. (1984) Herbicides as a possible cause of liver cancer. In:
    Westing, A.H., ed. Herbicides in war, the long-term ecological and
    human consequences, London, Taylor and Francis, pp. 119-121.

    VAN DEN BERG, M., OLIE, K., & HUTZINGER, O. (1983) Uptake and
    selective retention in rats of orally administered chlorinated dioxins
    and dibenzofurans from fly-ash and fly-ash extract. Chemosphere,
    12(4/5): 537-544.

    VAN DEN BERG, M., VAN GREEVENBROEK, M., & OLIE, K. (1986a)
    Bioavailability of PCDDs and PCDFs on fly ash after semi-chronic oral
    ingestion by the rat. Chemosphere, 15(4): 509-518.

    VAN DEN BERG, M., DE VROOM, E., & OLIE, K. (1986b) Bioavailability of
    PCDDs and PCDFs on fly ash after semi-chronic oral ingestion by guinea
    pig and syrian golden hamster. Chemosphere, 15(4): 519-533.

    VAN DEN BERG, M., VAN DER WIELEN, F.W.M., & OLIE, K. (1986) The
    presence of PCDDs and PCDFs in human breast milk from the Netherlands.
    Chemosphere, 15: 693-706.

    VAN LOGTEN, M.J., GUPTA, B.N., MCCONNELL, E.E., & MOORE, J.A. (1980)
    Role of the endocrine system in the action of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the thymus.
    Toxicology, 15: 135-144.

    VAN MILLER, J.P., MARLAR, R.J., & ALLEN, J.R. (1976) Tissue
    distribution and excretion of tritiated tetrachlorodibenzo-p-dioxin in
    non-human primates and rates. Food Cosmet. Toxicol., 14: 31-34.

    VAN MILLER, J.P., LALICH, J.J., & ALLEN, J.R. (1977) Increased
    incidence of neoplasms in rats exposed to low levels of
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Chemosphere, 61: 625-632.

    VECCHI, A., MANTOVANI, A., SIRONI, M., LUINI, W., CAIRO, M., &
    GARATTINI, S. (1980) Effect of acute exposure to
    2,3,7,8-tetrachlorodibenzo-p-dioxin on humoral antibody production in
    mice. Chem.-biol. Interact., 30: 337-342.

    VECCHI, A., SIRONI, M., CANEGRATI, M.A., RECCHIA, M., & GARATTINI, S.
    (1983) Immunosuppressive effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin
    in strains of mice with different susceptibility to induction of aryl
    hydrocarbon hydroxylase. Toxicol. appl. Pharmacol., 68: 434-441.

    VEERKAMP, W., WEVER, J., & HUTZINGER, O. (1981) The metabolism of some
    chlorinated dibenzofurans by rats. Chemosphere, 10: 397-403.

    VILLENUEVE, E.C., JENNINGS, R.W., BURSE, V.M., & KIMBROUGH, R.D.
    (1974) Evidence of chlorodibenzo-p-dioxin and chlorodibenzofuran in
    hexachlorobenzene. J. agric. food Chem., 22: 916-917.

    VINOPAL, J.H. & CASIDA, J.E. (1973) Metabolic stability of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin in mammalian liver microsomal
    systems and in living mice. Arch. environ. Contam. Toxicol., 1:
    122-132.

    VISWANATHAN, T.S. & KLEOPFER, R.D. (1986) The presence of
    hexachloroxanthene at Missouri dioxin sites, In: Rappe, C., Choudhary,
    G., & Keith, L., ed. Chlorinated dioxins and dibenzofurans in
    perspective, Chelsea, Michigan, Lewis Publishers, pp. 201-210.

    VOS, J.G. & BEEMS, R.B. (1971) Dermal toxicity studies of technical
    polychlorinated biphenyls and fractions thereof in rabbits. Toxicol.
    appl. Pharmacol., 19: 617-633.

    VOS, J.G. & KOEMAN, J.H. (1970) Comparative toxicologic study with
    polychlorinated biphenyls in chickens with special reference to
    porphyria, edema formation, liver necrosis, and tissue residues.
    Toxicol. appl. Pharmacol., 17: 656-668.

    VOS, J.G. & MOORE, J.A. (1974) Suppression of cellular immunity in
    rats and mice by maternal treatment with
    2,3,7,8-tetrachlorodibenzo-p-dioxin. Int. Arch. Allergy, 47:
    777-794.

    VOS, J.G., KOEMAN, J.H., VAN DER MAAS, H.L., TEN NOEVER DE BRAUW,
    M.C., & DE VOS, R.H. (1970) Identification and toxicological
    evaluation of chlorinated dibenzofuran and chlorinated napthalene in
    two commercial polychlorinated biphenyls. Food Cosmet. Toxicol.,
    8: 625-633.

    VOS, J.G., MOORE, J.A., & ZINKL, J.G. (1973) Effect of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on the immune system of laboratory
    animals. Environ. Health Perspect., 5: 149-162.

    VOS, J.G., MOORE, J.A., & ZINKL, J.G. (1974) Toxicity of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in C57 B1/6 mice.
    Toxicol. appl. Pharmacol., 29: 229-241.

    VOS, J.G., KREEFTENBERG, J.G., ENGEL, H.W.B., MINDERHOUD, A., & VAN
    NOORLE JANSEN, L.M. (1978a) Studies on
    2,3,7,8-tetra-chlorodibenzo-p-dioxin-induced immune suppression and
    decreased resistance to infection: Endotoxin hypersensitivity, serum
    zinc concentrations and effect of thymosin treatment. Toxicology,
    9: 75-86.

    VOS, J.G., KREEFTENBERG, J.G., & KATER, L. (1978b) Immune suppression
    by TCDD. In: Cattabeni, F., Cavallaro, A., & Galli, ed. Dioxin
    (TCDD), New York, John Wiley & Sons, Halsted Press Division, pp.
    163-175.

    WAHLE, P. (1914) [On two cases of chloracne.] Inaugural dissertation
    for the Doctorate in Medicine, Surgery and Obstetrics of the Medical
    Faculty, University of Leipzig (in German).

    WALDEN, R. & SCHILLER, C.M. (1985) Short communications. Comparative
    toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in four
    (sub)strains of adult male rats. Toxicol. appl. Pharmacol., 77:
    490-495.

    WARD, C.T. & MATSUMURA, F. (1978) Fate of
    2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD) in a model aquatic
    environment. Arch. environ. Contam. Toxicol., 7: 349-357.

    WAUER, (1918) [Occupational illness caused by chlorinated hydrocarbons
    (perna disease).] Zentralbl. Gewerbehyg., 6: 100-101 (in German).

    WEBER, H., & BIRNBAUM, L. S. (1985)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin (TCDD) and
    2,3,7,8-tetrachlorodibenzofuran (TCDF) in pregnant C57Bl/6N mice:
    Distribution to the embryo and excretion. Arch. Toxicol., 57:
    159-162.

    WEBER, H., POIGER, H., & SCHLATTER, Ch. (1982a) Acute oral toxicity of
    TCDD-metabolites in male guinea pigs. Toxicol. Lett., 14: 117-122.

    WEBER, H., POIGER, H., & SCHLATTER, Ch. (1982b) Fate of
    2,3,7,8-tetrachlorodibenzo-p-dioxin metabolites from dogs in rats.
    Xenobiotica, 12: 353-357.

    WEBER, H., LUZI, P., RESI, L., TANGANELLI, P., LOVATI, M.R., & POLI,
    A. (1983) Natural history of TCDD-induced liver lesions in rats as
    observed by transmission electron microscopy during a 32-week period
    after a single intraperitoneal injection. J. Toxicol. environ.
    Health, 12: 533-540.

    WEBER, H., LAMB, J.C., HARRIS, M.W., & MOORE, J.A. (1984)
    Teratogenicity of 2,3,7,8-tetrachlorodibenzofuran (TCDF) in mice.
    Toxicol. Lett., 20: 183-188.

    WEBER, H., HARRIS, M.W., HASEMAN, J.K., & BIRNBAUM, L.S. (1985)
    Teratogenic potency of TCDD, TCDF and TCDD-TCDF combinations in
    C57Bl/6N mice. Toxicol. Lett., 26: 159-167.

    WEBSTER, G.R.B., FRIESEN, K.J., SARNA, L.P., & MUIR, D.C.G. (1985)
    Environmental fate modelling of chlorodioxins: Determination of
    physical constants. Chemosphere, 14: 609-622.

    WEDEL, J., VON, HOLLA, W.A., & DENTON, J. (1943) Observations on the
    toxic effects resulting from exposure to chlorinated naphtalene and
    chlorinated phenyls with suggestions for prevention. Rubber Age,
    53: 419-426.

    WEISSBERG, J.B. & ZINKL, J.G. (1973) Effects of
    2,3,7,8-tetra-chlorodibenzo-p-dioxin upon hemostasis and hematologic
    function in the rat. Environ. Health Perspect., 5: 119-123.

    WESTING, A.H., (1984) Herbicides in war: past and present. In:
    Westing, A.H., ed. Herbicides in war, the long-term ecological and
    human consquences, London, Taylor and Francis.

    WHITE, K.L., Jr, LYSY, H.H., MCCAY, J.A., & ANDERSON, A.C. (1986)
    Modulation of serum complement levels following exposure to
    polychlorinated dibenzo-p-dioxins. Toxicol. appl. Pharmacol., 84:
    209-219.

    WHITLOCK, J.P. & GALEAZZI, D.R. (1984)
    2,3,7,8-Tetrachloro-dibenzo-p-dioxin receptors in wild type and
    variant mouse hepatoma cells. J. biol. Chem., 259: 980-985.

    WHO/EURO (1987) Dioxins and furans from municipal incinerators,
    Copenhagen, WHO Regional Office for Europe (Environmental Health
    Series 17).

    WIKLUND, K. & HOLM, L.-E. (1986) Soft tissue sarcoma risk in Swedish
    agricultural and forestry workers. J. Natl Cancer Inst., 76:
    229-234.

    WILLEY, J.C., SALADINO, A.J., OZANNE, C., LECHNER, J.F., & HARRIS,
    C.C. (1984) Acute effects of 12-O-tetradecanoyl-phorbol-13-acetate,
    teleocidin B, or 2,3,7,8-tetrachloro-dibenzo-p-dioxin on cultured
    normal human bronchial epithelial cells. Carcinogenesis, 5:
    209-215.

    WILLIAMS, D.T., CUNNINGHAM, H.M., & BLANCHFIELD, B.J. (1972)
    Distribution and excretion studies of octachlorodibenzo-p-dioxin in
    the rat. Bull. Environ. Contam. Toxicol., 7: 57-62.

    WIPF, H.K. & SCHMID, J. (1983) Seveso - An environmental assessment.
    In: Tucker, R.E., Young, A.L., & Gray, R., ed. Human and
    environmental risks of chlorinated dioxins and related compounds,
    New York, London, Plenum Press, pp. 255-276.

    WIPF, H.K., HOMBERGER, E., NEUNER, N., RANALDER, U.B., VETTER, W., &
    VUILLEUMIER, J.P. (1982) TCDD-levels in soil and plant samples from
    the Seveso area. In: Hutzinger, O., Frei, R.W., Merian, E., &
    Pocchiari, F., ed. Chlorinated dioxins and related compounds. Impact
    on the environment, Oxford, New York, Pergamon Press, pp. 115-127.

    WOLF, G. (1980) Vitamin A. In: Alfin-Slater, R.B. & Kritchevsky, D.,
    ed. Human nutrition, a comprehensive treatise. Part B. Nutrition and
    the adult, New York, London, Plenum Press, Vol. 3, pp. 97-203.

    WOODS, J.S. (1973) Studies of the effects of 2,3,7,8-tetra-
    chlorodibenzo-p-dioxin on mammalian hepatic delta-amino- levulinic
    acid synthetase. Environ. Health Perspect., 5: 221-225.

    WROBLEWSKI, V.J. & OLSON, J.R. (1985) Hepatic metabolism of
    2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the rat and guinea pig.
    Toxicol. appl. Pharmacol., 81: 231-240.

    WU, Y.C., LO, Y.C., KAO, H.Y., PAN, C.C., & LIN, R.Y. (1984)
    Cell-mediated immunity in patients with polychlorinated bi-phenyl
    poisoning. J. Formoson Med. Assoc., 83: 419-429.

    YAMAGISHI, T., MIYAZAKI, T., AKIYAMA, K., MORITA, M., NAKAGAWA, J.,
    HORII, S., & KANEKO, S. (1981) Polychlorinated dibenzo-p-dioxins and
    dibenzofurans in commercial diphenyl ether herbicides, and in fresh
    water fish collected from the application area. Chemosphere, 10:
    1137-1144.

    YANG, K.H., CROFT, W.A., & PETERSON, R.E. (1977) Effects of
    2,3,7,8-tetrachlorodibenzo-p-dioxin on plasma disappearance and
    biliary excretion of foreign compounds in rats. Toxicol. appl.