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.

    First draft prepared by Dr S. Dobson, Institute of Terrestrial
    Ecology, United Kingdom, and Dr G.J. van Esch, Bilthoven, The

    World Health Organization
    Geneva, 1993

        The International Programme on Chemical Safety (IPCS) is a joint
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    with chemical accidents, coordination of laboratory testing and
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    WHO Library Cataloguing in Publication Data

    Polychlorinated Biphenyls and Terphenyls. -- 2nd ed.

    (Environmental health criteria; 140)

    1.Environmental exposure 2.Environmental pollutants 3.Polychlorinated
    biphenyls -- adverse effects 4.Polychlorinated biphenyls -- toxicity
    5.Polychloroterphenyl compounds -- adverse effects
    6.Polychloroterphenyl compounds -- toxicity I.Series

    ISBN 92 4 157140 3 (NLM Classification: QV 633)
    ISSN 0250-863X

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         1.1    Summary and evaluation
                1.1.1    Introduction
                1.1.2    Identity, physical, and chemical properties
                1.1.3    Analytical methods
                1.1.4    Production and uses
                1.1.5    Environmental transport, distribution, and transformation
                1.1.6    Environmental levels and human exposure
                1.1.7    Kinetics and metabolism
                1.1.8    Effects on organisms in the environment
                   Laboratory studies
                   Field studies
                1.1.9    Effects on experimental animals and  in vitro systems
                   Single exposure
                   Short-term exposure
                1.1.10   Reproduction, embryotoxicity, and teratogenicity
                1.1.11   Mutagenicity
                1.1.12   Carcinogenicity
                1.1.13   Special studies
                1.1.14   Factors modifying toxicity, mode of action
                1.1.15   Effects on humans
         1.2    Conclusions
                1.2.1    Distribution
                1.2.2    Effects on experimental animals
                1.2.3    Effects on humans
                1.2.4    Effects on the environment
         1.3    Recommendations

         2.1    Identity
                2.1.1    Chemical formula and structure
                2.1.2    Relative molecular mass
                2.1.3    Common name
                2.1.4    Chemical composition
                2.1.5    Technical product
                2.1.6    Purity and impurities
         2.2    Physical and chemical properties
                2.2.1    Log  n-octanol/water partition coefficient
                2.2.2    Conversion factors

         2.3    Analytical methods
                2.3.1    Sampling strategy and sampling methods
                   Extraction procedures
                   Sample clean-up
                2.3.2    Separation and identification
                   Chromatographic separation
                   Gas-liquid chromatography
                2.3.3    Quantification
                2.3.4    Accuracy of PCB determinations
                2.3.5    Confirmation
                2.3.6    Detection limits
         2.4    Codex questionnaire on analytical methods
                2.4.1    Interpretation and comparability of data
         2.5    Activities of the WHO Regional Office for Europe
         2.6    Appraisal

         3.1    Natural occurrence
         3.2    Man-made sources
                3.2.1    Production levels and processes, uses
                   World production figures
                   Manufacturing processes
                3.2.2    Uses
                   Completely closed systems
                   Nominally closed systems
                   Open-ended applications
                   Contamination of other compounds
                3.2.3    Loss into the environment
                   Routes of environmental pollution
                   Release of PCBs into the atmosphere
                   Leakage and disposal of PCBs in industry
                3.2.4    Thermal decomposition of PCBs

         4.1    Transport and distribution between media
                4.1.1    Transport in air
                   Dry deposition
                   Precipitation deposition
                4.1.2    Transport in soil
                4.1.3    Transport in water
                4.1.4    Transport between media
         4.2    Biotransformation
                4.2.1    Biodegradation
                4.2.2    Biodegradation; individual congeners

                4.2.3    Photodegradation
                4.2.4    Bioaccumulation, distribution in organisms, and elimination
                   Aquatic invertebrates
                4.2.5    Appraisal

         5.1    Levels in the environment
                5.1.1    Air
                   Rain and snow
                   Natural gas
                5.1.2    Water
                5.1.3    Soil
                5.1.4    Aquatic and terrestrial organisms
                   Effect of dredging-contaminated sediment on organisms
                   Relationship to lipid content of organisms
                   Residues in different trophic levels and effects of diets
                   Effects of age, sex, and reproductive status on uptake and elimination
                   Time trends in residues
                   Seasonal patterns in residues
                5.1.5    Appraisal
         5.2    Levels in animal feed
         5.3    Levels in human food
                5.3.1    General
                5.3.2    Drinking-water
                5.3.3    Dairy products
                5.3.4    Fish and shellfish
                5.3.5    Influence of food processing
                5.3.6    Food contamination by packaging materials
                5.3.7    Appraisal
         5.4    General population exposure
                5.4.1    Air
                5.4.2    Drinking-water
                5.4.3    Intake by infants through mother's milk
                5.4.4    Infant and toddler total diet
                5.4.5    Total intake by adults via food
                5.4.6    Total diet/market-basket studies
                5.4.7    Total intake of major congeners by adults via food
                5.4.8    Time trends in different matrices

         5.5    Concentrations in the body tissues of the general population
                5.5.1    Adipose tissue
                   PCBs in the fetus
                   Congeners in adipose tissue
                5.5.2    Blood of the general population
                5.5.3    Human milk
                   Major PCB congeners in human milk
                   Factors that influence the intake of PCBs with milk
                5.5.4    Other tissues
         5.6    Accidental exposures (Yusho and Yu-Cheng)
         5.7    Occupational exposure
                5.7.1    Accidental exposure
                5.7.2    Occupational exposure during manufacture and use
                   Adipose tissue

         6.1    Absorption
                6.1.1    Inhalation
                6.1.2    Dermal
                6.1.3    Oral
         6.2    Distribution
                6.2.1    Inhalation (rat)
                6.2.2    Oral (rat)
                6.2.3    Oral (monkey)
                6.2.4    Oral (humans)
                6.2.5    Individual congeners of PCBs
                6.2.6    Appraisal
         6.3    Placental transport
                6.3.1    Laboratory animals
                6.3.2    Wildlife
                6.3.3    Humans
         6.4    Excretion and elimination
                6.4.1    Following oral dosing
                6.4.2    Following parenteral dosing
                6.4.3    Humans
                6.4.4    Elimination via milk (animals)
                   Elimination via breast milk
         6.5    Metabolic transformation
                6.5.1    PCBs
                6.5.2    Dichlorobiphenyls
                6.5.3    Tetrachlorobiphenyls
                6.5.4    Hexachlorobiphenyls and higher chlorinated compounds
                6.5.5    Retention and turnover
                6.5.6    Appraisal

         7.1    Toxicity for microorganisms
                7.1.1    Freshwater microorganisms
                7.1.2    Marine and estuarine microorganisms
                7.1.3    Soil microorganisms
                7.1.4    Plankton communities
                7.1.5    Interactions with other chemicals
                7.1.6    Tolerance
         7.2    Toxicity for aquatic organisms
                7.2.1    Aquatic plants
                7.2.2    Aquatic invertebrates
                   Short- and long-term toxicity
                   Response to temperature and salinity
                   Population structure
                   Interactions with other chemicals
                7.2.3    Fish
                   Short- and long-term toxicity
                   Effects on developmental stages and reproduction
                   Physiological and biochemical effects
                   Behavioural effects
                   Interactions with other chemicals
                7.2.4    Amphibians
                7.2.5    Aquatic mammals
         7.3    Toxicity for terrestrial organisms
                7.3.1    Plants
                7.3.2    Terrestrial invertebrates
                7.3.3    Birds
                   Short-term toxicity
                   Egg production
                   Hatchability and embryotoxicity
                   Eggshell thinning
                   Effects on the male
                   The effects of stress
                   Physiological, biochemical, and behavioural effects
                   Interactive effects with other chemicals
                7.3.4    Terrestrial mammals
                   Short-term toxicity
                   Reproductive effects
                   Physiological effects

         7.4    Effects on organisms in the field
                7.4.1    Plants
                7.4.2    Fish
                7.4.3    Birds
                7.4.4    Mammals

         8.1    Single exposures
                8.1.1    Oral
                8.1.2    Inhalation
                8.1.3    Dermal
                8.1.4    Other routes
         8.2    Short-term exposures
                8.2.1    Oral
                   Individual congeners
                8.2.2    Intraperitoneal: reconstituted PCB mixtures
                8.2.3    Dermal exposure
                8.2.4    Appraisal
         8.3    Skin and eye irritation, sensitization
         8.4    Reproduction, embryotoxicity, and teratogenicity
                8.4.1    Reproduction and embryotoxicity
                8.4.2    Teratogenicity
                   Aroclors (oral)
                   Aroclors (subcutaneous)
                   Individual congeners (oral)
                8.4.3    Appraisal
                8.4.4    Mutagenicity and related end-points
                   DNA damage
                   Mutagenicity tests
                   Cell transformation
                   Cell to cell communication
                   Cell division parameters
         8.5    Carcinogenicity
                8.5.1    Long-term toxicity/carcinogenicity
                8.5.2    Tumour promotion/anticarcinogenic effects
                8.5.3    Initiation, promotion, and other special studies on individual congeners
                8.5.4    Skin carcinogenicity
                8.5.5    Appraisal
         8.6    Special studies: target-organ effects
                8.6.1    Liver
                   PCB mixtures
                   Individual congeners

                8.6.2    Enzyme induction
                   Effects on liver enzymes of PCBs
                   Effects on liver enzymes of "biologically filtered" PCB mixtures
                   Effects of individual congeners on liver enzymes
                8.6.3    Effects on vitamins and mineral metabolism
                   Effects of PCB mixtures
                   Effects of individual congeners
                8.6.4    Effects on the gastrointestinal tract
                8.6.5    Effects on lipid metabolism
                   Effects of PCB mixtures
                   Effects of individual congeners
                8.6.6    Effects on porphyrin metabolism
                   Effects of PCB mixtures
                   Effects of individual congeners
                8.6.7    Effects on the endocrine system
                   Effects of PCB mixtures
                   Effects of individual congeners
                8.6.8    Immunotoxicity
                   Effects of PCB mixtures
                   Effects of individual congeners
                8.6.9    Neurotoxic effects
                8.6.10   Skin effects
                8.6.11   Effects on the lung
                8.6.12   Miscellaneous
         8.7    Factors modifying toxicity; mode of action
                8.7.1    Factors modifying toxicity
                8.7.2    Mechanisms of toxicity
                8.7.3    Toxicity of impurities in commercial PCBs

         9.1    General population exposure
                9.1.1    Acute effects - poisoning incidents
                9.1.2    Effects of short- and long-term exposure
                   Yusho and Yu-Cheng accidents
                   Effects of PCBs on babies and infants
                9.1.3    Appraisal
         9.2    Occupational exposure
                9.2.1    Acute toxicity - poisoning incidents
                   Acute dermal effects
                9.2.2    Effects of short- and long-term exposure
                9.2.3    Appraisal

                9.2.4    Special studies (target organ effects)
                   Effects on the liver
                   Effects on the respiratory system
                   Blood pressure
                9.2.5    Mortality studies
                9.2.6    Appraisal




         1.1    Identity
         1.2    Physical and chemical properties
         1.3    Analytical methods



         4.1    Residues in the environment
         4.2    Residues in food
         4.3    Concentrations in adipose tissue
         4.4    Concentrations in blood

         5.1    Absorption
         5.2    Distribution
         5.3    Biotransformation

         6.1    Marine and estuarine organisms
         6.2    Terrestrial invertebrates
         6.3    Birds

         7.1    Single oral exposures
         7.2    Short-term oral exposures
                7.2.1    Rat
                7.2.2    Monkey
         7.3    Teratogenicity
         7.4    Carcinogenicity
         7.5    Miscellaneous effects


    ANNEX 1






    Dr L.A. Albert, Consultores Ambientales Asociados, Xalapa, Veracruz,

    Professor U.G. Ahlborg, Institute of Environmental Medicine,
    Karolinska Institute, Stockholm, Sweden

    Dr V. Benes, Department of Toxicology and Reference Laboratory,
    Institute of Hygiene and Epidemiology, Prague, Czechoslovakia

    Dr S. Dobson, Institute of Terrestrial Ecology, Monks Wood
    Experimental Station, Abbots Ripton, Huntingdon, United Kingdom

    Dr Yuzo Hayashi, Division of Pathology, National Institute of Hygienic
    Sciences, Tokyo, Japan

    Dr T. Lakhanisky, Division of Toxicology, Institute of Hygiene and
    Epidemiology, Brussels, Belgium

    Dr J. McKinney, US Environmental Protection Agency, Research Triangle
    Park, North Carolina, USA

    Dr Pang Ying Fa, Chinese Academy of Preventive Medicine, Beijing,

    Dr T. Vermeire, National Institute of Public Health and Environmental
    Protection, Bilthoven, Netherlands  (Co-Rapporteur)

    Dr E. Yrjnheikki, Regional Institute of Occupational Health, Oulu,


    Dr M. Martens (Representative from ECETOC), Monsanto Services
    International, Brussels, Belgium

    Mrs H. B. Sundmark (Representative from ECETOC), Norsk Hydro a.s.
    Porsgrunn, Research Centre, Porsgrunn, Norway


    Dr G.J. van Esch, Bilthoven, Netherlands  (Co-Rapporteur and

    Dr M. Kogevinas, Unit of Analytical Epidemiology, International Agency
    for Research on Cancer (IARC), Lyon, France


    Every effort has been made to present information in the criteria
    monographs as accurately as possible without unduly delaying their
    publication. In the interest of all users of the environmental health
    criteria monographs, readers are kindly requested to communicate any
    errors that may have occurred to the Director 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).


    A WHO Task Group on Environmental Health Criteria for PCBs and PCTs
    met in Brussels from 28 May to 1 June 1990. The meeting was convened
    in the Institute of Hygiene and Epidemiology in Brussels and sponsored
    by the Belgian Ministry of Health. Mrs A.-M. Sacr-Bestin of the
    Ministry opened the meeting and welcomed the participants on behalf of
    the host country. Dr G.J. van Esch welcomed the participants on behalf
    of the Heads of the three IPCS cooperating organizations
    (UNEP/ILO/WHO). The Group reviewed and revised the draft Environmental
    Health Criteria monograph and the companion Health and Safety Guide
    and made an evaluation of the risks for human health and the
    environment from exposure to PCBs and PCTs.

    The first draft of the EHC monograph was prepared by Dr S. Dobson
    (environmental aspects) and Dr G.J. van Esch (other sections) and was
    based on contributions from several authors and countries. It was
    prepared in close cooperation with the WHO Regional Office for Europe,
    in Copenhagen.

    The second draft was prepared by Dr G.J. van Esch, incorporating
    comments received following the circulation of the first draft to the
    IPCS contact points for Environmental Health Criteria monographs.
    Dr K. Jager, Central Unit, IPCS, was responsible for the scientific
    content of the final monograph and Mrs M.O. Head, Oxford, for the

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


    The commercial production of the polychlorinated biphenyls (PCBs)
    began in 1930, and, during the 1930s, cases of poisoning were reported
    among men engaged in their manufacture. The nature of this
    occupational disease was characterized by a skin affection with
    acneiform eruptions; occasionally the liver was involved, in some
    cases with fatal consequences. Subsequent safety precautions appear
    largely to have prevented further outbreaks of this disease in
    connection with the manufacture of PCBs, but, since 1953, cases have
    been reported in Japanese factories manufacturing condensers.

    The distribution of PCBs in the environment was not recognized until
    Jensen started an investigation in 1964 to ascertain the origins of
    unknown peaks, observed during the gas-liquid chromatographic
    separation of organochlorine pesticides from wildlife samples. In
    1966, he and his colleagues succeeded in attributing these to the
    presence of PCBs. Since then, investigations in many parts of the
    world have revealed the widespread distribution of PCBs in
    environmental samples.

    The serious outbreaks of poisoning in humans and in domestic animals
    from the ingestion of food, accidentally contaminated with PCBs, have
    stimulated investigations into the toxic effects of PCBs on animals
    and on nutritional food chains. This has resulted in the limitation of
    the commercial exploitation of PCBs and polychlorinated terphenyls
    (PCTs), and in regulations to limit the residues in human and animal

    In recent years, many industrial nations have taken steps to control
    the flow of PCBs into the environment. PCBs and PCB-containing
    formulations are restricted (an exception is sometimes made for mono-
    and dichloro-PCBs) for most uses. Now they are almost entirely
    restricted to use in closed systems, such as isolating oils in
    transformers, capacitors, and other electrical systems, and as a heat
    transfer medium and hydraulic liquid. The most influential forces
    leading to these restrictions have probably been the 1973 and 1987
    decision-recommendations from the Organisation for Economic
    Co-operation and Development (OECD).

    The environmental impact of the PCBs and PCTs has been discussed at a
    number of regional and international meetings and has been the subject
    of several reviews, including: ATSDR (1989), DFG (1988), IARC (1978),
    IRPTC (1988), Kimbrough (1987), Lorenz & Neumeier (1983a,b), NIOSH
    (1987), NTIS (1972), OECD (1982), Slorach & Vaz (1983), WHO (1985a,b,
    1986a,b) & WHO/EUR (1987).

    In 1976, the World Health Organization published Environmental Health
    Criteria 2: Polychlorinated biphenyls (PCBs) and terphenyls (PCTs)
    (WHO, 1976), discussing and evaluating the data then available on
    exposure levels and the effects of PCBs and PCTs on human beings, and,
    to a lesser extent, on the environment.

    Since then, a wealth of new information has become available.

    The IPCS decided to update the above-mentioned EHC and also to produce
    a Health and Safety Guide (HSG) and to do this in close coordination
    with the WHO Regional Office for Europe, which prepared "PCBs, PCDDs
    and PCDFs, prevention and control of accidental and environmental
    exposures" as No. 23 of their Environmental Health Series (WHO/EURO,
    1987). This publication includes a set of guidelines to assist Member
    States in the development of strategies to reduce the probability of
    accidents involving the environmental release of PCBs, PCDDS, and
    PCDFs and also the severity of their hazardous effects, should such
    accidents occur. In particular, it is intended to guide occupational
    safety and health personnel and other staff, in workplaces and
    environments where PCBs and/or PCB-containing equipment are in use, to
    develop adequate safety measures, contingency planning, effective and
    relevant accident response, and appropriate rehabilitation.

    Within the scope of the present EHC on PCBs and PCTs, the PCDDs and
    PCDFs have been mentioned where relevant. Full discussion of these
    compounds and evaluation, however, can be found in the IPCS EHC 88:
    Polychlorinated dibenzo- para-dioxins and dibenzofurans (WHO, 1989).


    1.1   Summary and evaluation

    1.1.1  Introduction

    Polychlorinated biphenyls (PCBs) were discovered before the turn of
    the century and their usefulness for industry, because of their
    physical properties, was recognized early. The PCBs have been used
    commercially, since 1930, as dielectric and heat-exchange fluids and
    in a variety of other applications. They have become widely
    distributed in the environment throughout the world, and are
    persistent and accumulate in food webs. Human exposure to PCBs has
    resulted largely from the consumption of contaminated food, but also
    from inhalation and skin absorption in work environments. PCBs
    accumulate in the fatty tissues of humans and other animals and have
    caused toxic effects in both, particularly if repeated exposure
    occurs. The skin and liver are the major sites of pathology, but the
    gastrointestinal tract, the immune system, and the nervous system are
    also targets. Polychlorinated dibenzofurans (PCDFs), which are
    contaminants in commercial PCB mixtures, contribute significantly to
    their toxicity. The results of studies on rodents suggest that some
    PCB congeners may be carcinogenic and that they can promote the
    carcinogenicity of other chemicals.

    It is clear from available data on polychlorinated biphenyls (PCBs)
    and polychlorinated terphenyls (PCTs) that, in an ideal situation, it
    would be preferable not to have these compounds in food at any level.
    However, it is equally clear that the reduction of PCBs or PCTs
    exposure from food sources to "zero" or to a level approaching zero,
    would mean the elimination (prohibition of the consumption) of large
    amounts of important food items, such as fish, but more importantly
    breast milk. National and international scientific committees have to
    decide where the proper balance lies between providing an adequate
    degree of public health protection and avoiding excessive losses of

    No levels of PCBs or PCTs exposure that can provide an absolute
    assurance of safety can be identified on the basis of the available

    1.1.2  Identity, physical, and chemical properties

    PCBs are mixtures of aromatic chemicals, manufactured by the
    chlorination of biphenyl in the presence of a suitable catalyst. The
    chemical formula of PCBs can be presented as C12H10-nCln, where n is
    a number of chlorine atoms within the range of 1-10.

    Theoretically, 209 congeners are possible, but only about 130
    congeners are likely to occur in commercial products. In addition,
    PCBs may contain polychlorinated dibenzofurans (PCDFs) and chlorinated
    quarterphenyls as impurities. These impurities are relatively stable
    and resistant to chemical reactions, under normal conditions. All
    congeners of PCBs are lipophilic and have a very low water solubility.
    As a result, they easily enter the food chain and accumulate in fatty

    Commercial PCB mixtures contain PCDFs at levels ranging from a few
    mg/kg up to 40 mg/kg. Polychlorinated dibenzo- p-dioxins (PCDDs), are
    not found in commercial PCBs. However, when PCBs are mixed with other
    chlorinated compounds, such as the chloro-benzenes used in
    transformers, PCDDs can be found in the case of accidental fires and
    during incineration.

    Commercial PCB mixtures are light yellow or dark yellow in colour.
    They do not crystallize, even at low temperatures, but turn into solid
    resins. PCBs are, in practice, fire resistant, with rather high flash
    points. They form vapours heavier than air, but they do not form any
    explosive mixtures with air. They have very low electrical
    conductivity, rather high thermal conductivity, and extremely high
    resistance to thermal break-down. PCBs are chemically very stable
    under normal conditions; however, when heated, other toxic compounds,
    such as PCDFs, can be produced.

    1.1.3  Analytical methods

    In 1966, the discovery of PCBs in environmental samples raised
    interest in the analysis of these compounds and their toxicity for
    human beings and their environment.

    Because of differences in the analytical methodology used, existing
    data are not directly comparable; nevertheless, they can be used for
    the establishment of control and preventive measures and for the
    preliminary assessment of health and environmental risks associated
    with these chemicals.

    PCBs have been determined using gas chromatography (GC) techniques
    with electron capture detection, often using packed columns, though
    more sophisticated methods, such as capillary column and GC coupled
    with mass-spectrometry (GC-MS), have been used in recent studies to
    identify the individual congeners, to improve the comparability of the
    analytical data from different sources, and to establish a basis for
    toxicity assessment.

    An extensive quality assurance programme is required for these
    analyses and intercalibration studies have been implemented and
    recommended. The quality and utility of the analytical data depend
    critically on the validity of the sample and the adequacy of the
    sampling. Furthermore, it is essential to have a planned and well
    documented sampling programme; a detailed sampling procedure is
    described in WHO/EURO (1987).

    1.1.4  Production and uses

    The commercial production of the PCBs began in 1930. They have been
    widely used in electrical equipment, and smaller volumes of PCBs are
    used as fire-resistant liquid in nominally closed systems.

    By the end of 1980, the total world production of PCBs was in excess
    of 1 million tonnes and, since then, production has continued in some
    countries. Despite increasing withdrawal of the use, and restrictions
    on the production, of PCBs, very large amounts of these compounds
    continue to be present in the environment, either in use or as waste.

    In recent years, many industrialized countries have taken steps to
    control and restrict the flow of PCBs into the environment. The most
    influential force leading to these restrictions has probably been a
    1973 recommendation from the Organisation for Economic Co-operation
    and Development (OECD) (WHO, 1976; IARC, 1978; OECD, 1982). Since
    then, the 24 OECD member countries have restricted the manufacture,
    sales, importation, exportation, and use of PCBs, as well as
    establishing a labelling system for these compounds.

    Current sources of PCB release include volatilization from landfills
    containing transformer, capacitor, and other PCB-wastes, sewage
    sludge, spills, and dredge spoils, and improper (or illegal) disposal
    to open areas. Pollution may occur during the incineration of
    industrial and municipal waste. Most municipal incinerators are not
    effective in destroying PCBs. Explosions or overheating of
    transformers and capacitors may release significant amounts of PCBs
    into the local environment.

    PCBs can be converted to PCDFs under pyrolytic conditions. The highest
    yield of PCDFs under laboratory conditions was obtained at a
    temperature between 550 and 700C. Thus, the uncontrolled burning of
    PCBs can be an important source of hazardous PCDFs. It is therefore
    recommended that destruction of PCB-contaminated waste should be
    carefully controlled, especially with regard to the burning
    temperature (above 1000C), residence time, and turbulence.

    1.1.5  Environmental transport, distribution, and transformation

    In the atmosphere, PCBs exist primarily in the vapour phase; the
    tendency to adsorb on particulates increases with the degree of
    chlorination. The virtually universal distribution of PCBs suggests
    transport in air.

    At present, the major source of PCB exposure in the general
    environment appears to be the redistribution of PCBs, previously
    introduced into the environment. This redistribution involves
    volatilization from soil and water into the atmosphere with subsequent
    transport in air and removal from the atmosphere via wet/dry
    deposition (of PCBs bound to particulates) and then re-volatilization.
    Concentrations of PCBs in precipitation range from 0.001 to
    0.25 g/litre. Since the volatilization and degradation rates of PCBs
    vary between congeners, this redistribution leads to an alteration in
    the composition of PCB mixtures in the environment.

    In water, PCBs are adsorbed on sediments and other organic matter;
    experimental and monitoring data have shown that PCB concentrations in
    sediment and suspended matter are higher than those in associated
    water columns. Strong adsorption on sediment, especially in the case
    of the higher chlorinated PCBs, decreases the rate of volatilization.
    On the basis of their water solubilities and  n-octanol-water
    partition coefficients, the lower chlorinated PCB congeners will sorb
    less strongly than the higher chlorinated isomers. Although adsorption
    can immobilize PCBs for relatively long periods in the aquatic
    environment, desorption into the water column has been shown to occur
    by both abiotic and biotic routes. The substantial quantities of PCBs
    in aquatic sediments can therefore act as both an environmental sink
    and a reservoir of PCBs for organisms. Most of the environmental load
    of PCBs has been estimated to be in aquatic sediment.

    The low solubility and the strong adsorption of PCBs on soil particles
    limits leaching in soil; lower chlorinated PCBs will tend to leach
    more than the highly chlorinated PCBs.

    Degradation of PCBs in the environment is dependent on the degree of
    chlorination of the biphenyl. In general, persistence of PCB congeners
    increases as the degree of chlorination increases. In the atmosphere,
    the vapour phase reaction of PCBs with hydroxyl radicals (which are
    photochemically formed by sunlight) may be the dominant transformation
    process. Estimated half-lives for this reaction in the atmosphere
    range from about 10 days for a monochlorobiphenyl to 1.5 years for a

    In the aquatic environment, hydrolysis and oxidation do not
    significantly degrade PCBs. Photolysis appears to be the only viable
    abiotic degradation process in water; however, available experimental
    data are not sufficient to determine its rate or importance in the

    Microorganisms degrade mono-, di-, and trichlorinated biphenyls
    relatively rapidly and tetrachlorobiphenyls slowly, whilst higher
    chlorinated biphenyls are resistant to biodegradation. Chlorine
    substitution positions on the biphenyl ring appear to be important in
    determining the biodegradation rate. PCBs containing chlorine atoms in
    the  para positions are preferentially biodegraded. Higher
    chlorinated congeners are biotransformed anaerobically, by a reductive
    dechlorination, to lower chlorinated PCBs, which may then be
    biodegradable by aerobic processes.

    Several factors determine the degree of bioaccumulation in adipose
    tissues: the duration and level of exposure, the chemical structure of
    the compound, and the position and pattern of substitution. In
    general, the higher chlorinated congeners are accumulated more

    Experimentally determined bioconcentration factors of various PCBs in
    aquatic species (fish, shrimp, oyster) range from 200 up to 70 000 or
    more. In the open ocean, there is bioaccumulation of PCBs in higher
    trophic levels with an increased proportion of higher chlorinated
    biphenyls in higher ranking predators.

    Transfer of PCBs from soil to vegetation takes place mainly by
    adsorption on the external surfaces of terrestrial plants; little
    translocation takes place.

    1.1.6  Environmental levels and human exposure

    Because of their high persistence, and their other physical and
    chemical properties, PCBs are present in the environment all over the

    Globally, PCBs are found in air concentrations of 0.002 up to
    15 ng/m3. In industrial areas, levels are higher (up to g/m3). In
    rain water and snow, PCBs are found in the range of nd (1 ng)-
    250 ng/litre.

    Under occupational conditions, the levels in the air may be much
    higher. Under certain conditions, for instance, in the manufacturing
    of transformers or capacitors, levels of up to 1000 g/m3 have been
    observed. In acute emergencies, concentrations of up to 16 mg/m3 have
    been measured. In case of fires and/or explosions, soot may be
    produced that contains high levels of PCBs. Levels of 8000 mg PCBs/kg
    soot have been found. In the latter situation, PCDFs will also be
    present. Polychlorinated dioxins (PCDDs) will be found in accidents
    with transformers containing chlorinated benzenes, as well as PCBs.

    In these emergency situations, ingestion, skin contamination, or
    inhalation of soot particles may occur and result in serious exposure
    of personnel. However, the exposure of the general population via air
    will be very low.

    Surface water may be contaminated by PCBs from atmospheric fallout,
    from direct emissions from point sources, or from waste disposal.
    Under certain conditions, levels of up to 100-500 ng/litre water have
    been measured. In the oceans, levels of 0.05-0.6 ng/litre have been

    In non-contaminated areas, drinking-water contains less than 1 ng
    PCBs/litre, but levels of up to 5 ng/litre have been reported. Soil
    and sediments in different areas and depending on local conditions,
    contain levels of PCBs ranging from <0.01 up to 2.0 mg/kg. In
    polluted areas, the levels have been much higher, i.e., up to
    500 mg/kg.

    In past years, many thousands of samples of different foodstuffs have
    been analysed in several countries for contaminants including PCBs.
    Most samples have been taken from individual food items, especially
    fish and other foods of animal origin, such as meat and milk. Human
    food has become contaminated with PCBs by 3 main routes:

     (a) uptake from the environment by fish, birds, livestock (via
    food-chains), and crops;

     (b) migration from packaging materials into food (mainly below
    1 mg/kg, but, in some cases, up to 10 mg/kg);

     (c) direct contamination of food or animal feed by an industrial

    The levels for the most important PCB-containing food items were:
    animal fat, 20-240 g/kg; cow's milk, 5-200 g/kg; butter,
    30-80 g/kg; fish, 10-500 g/kg, on a fat basis. Certain fish species
    (eel) or fish products (fish liver and fish oils) contain much higher
    levels, up to 10 mg/kg. Vegetables, cereals, fruits, and a number of
    other products contained levels of <10 g/kg. The major foods in
    which contamination with PCBs needs consideration are fish, shellfish,
    meat, milk, and other dairy products. Median levels in fish, reported
    in various countries, are of the order of 100 g/kg (on a fat basis).
    When comparisons have been made, it appears that the levels of PCBs in
    fish are slowly decreasing.

    PCBs concentrate in human adipose tissue and breast milk. The
    concentrations of PCBs in the different organs and tissues depend on
    their lipid contents, with the exception of the brain. PCB residues in
    the adipose tissue of the general population in industrialized
    countries range from less than 1 up to 5 mg/kg, on a fat basis.

    The average concentrations of total PCBs in human milk fat are in the
    range of 0.5-1.5 mg/kg fat, depending on the donor's residence,
    life-style, and the analytical methods used. Women who live in
    heavily-industrialized, urban areas, or who consume a lot of fish,
    especially from heavily-contaminated waters, may have higher PCB
    concentrations in their breast milk.

    The composition of most PCB extracts from environmental samples does
    not resemble that of the commercial PCB mixtures. It has also been
    shown, using high-resolution gas chromatography analysis, that the
    congener composition and the relative concentrations of the individual
    components in adipose tissues and breast milk differ markedly from
    those in the commercial PCBs. The GC patterns of PCBs in human adipose
    tissue and breast milk contain relatively high concentrations of
    mainly the higher chlorinated PCBs, such as: 2,4,5,3',4'-pentachloro
    biphenyl; 2,4,5,2',4',5'-hexachlorobiphenyl, and 2,3,4,2',4',5'-
    hexachlorobiphenyl; 2,3,4,5,2',4',5'-hepta- and 2,3,4,5,2',3',4'-
    heptachlorobiphenyl. A few other PCB congeners are present in
    much lower quantities, such as the most toxic, coplanar PCBs:
    3,4,3',4'-tetra-, 3,4,5,3',4'-penta-, and 3,4,5,3',4',5'-

    It has been calculated that the daily intake of PCBs by infants from
    breast milk, is of the order of 4.2 g/kg body weight (5.2 g/100 Kcal
    consumed) (WHO/EURO, 1988). The average total of ingested PCBs from
    breast milk, during the first 6 months of life, is 4.5 mg compared
    with the calculated intake of 357 mg of PCBs over the subsequent
    life-time (0.2 g/kg per day from the diet of a 70-kg person over a
    70-year life-time). Therefore, the nursing period contributes about
    1.3% of the life-time intake, which is not large, in the light of the
    benefits of breast-feeding (WHO/EURO, 1988).

    On the basis of the evaluated background data, for adults the average
    dietary intake of PCBs amounts to a maximum of 100 g per week, or
    approximately 14 g/person per day. For a 70-kg person, this is an
    intake equivalent to a maximum of 0.2 g/kg body weight per day
    (WHO/EURO, 1988).

    1.1.7  Kinetics and metabolism

    Animal studies have been reported involving mainly oral, inhalation,
    and dermal exposures to both PCB mixtures and individual congeners. In
    general, PCBs appear to be rapidly absorbed, particularly by the
    gastrointestinal tract after oral exposure. It is clear that
    absorption does occur in humans, but information on the rates of human
    absorption of PCBs is limited.

    From the available studies, the data on the distribution of PCBs,
    suggest a biphasic kinetic process with rapid clearance from the blood
    and accumulation in the liver and the adipose tissue of various
    organs. There is also evidence of placental transport, fetal
    accumulation, and distribution to milk. In some human studies, the
    skin contained a high concentration of PCBs, but the concentration in
    the brain was lower than that expected on the basis of the lipid

    Mobilization of PCBs from fat appears to depend largely on the rates
    of metabolism of the individual PCB congeners. Excretion depends on
    the metabolism of PCBs to more polar compounds, such as phenols,
    conjugates of thiol compounds, and other water-soluble derivatives.

    Metabolic pathways include hydroxylation, and conjugation with thiols
    and other water-soluble derivatives, some of which can involve
    reactive intermediates, such as the arene oxides. Rates of metabolism
    have been shown to depend on the PCB structure and reflect both the
    degree and position of chlorine substituents. The polar metabolites of
    the more highly chlorinated PCBs appear to be eliminated primarily in
    the faeces, but excretion in the urine can also be significant. An
    important elimination route, is via (breast) milk. Certain PCB
    congeners can also be eliminated via hair.

    The available kinetic studies indicate that there is a wide divergence
    in biological half-life among the individual congeners and this can
    reflect differences in structure-dependent metabolism, tissue
    affinities, and other factors affecting mobilization from storage
    sites. Persistence in tissues is not always correlated with high
    toxicity, and differences in toxicity between PCB congeners may be
    associated with specific metabolites and/or their intermediates.

    1.1.8  Effects on organisms in the environment

    PCBs are universal, environmental contaminants and are present in most
    environmental compartments, abiotic and biotic, throughout the world.
    Since many countries have controlled both use and release, new input
    into the environment is on a reduced scale compared with the past.
    However, the available evidence suggests that the cycling of PCBs is
    causing a gradual redistribution of some congeners towards the marine
    environment. There is a trend for the highest chlorinated congeners to
    accumulate preferentially. While much of the PCB is adsorbed on to
    particulates in sediment, it is still bioavailable to organisms and
    will continue to be accumulated in higher trophic levels.  Laboratory studies

    Effects of PCB mixtures on microorganisms are highly variable with
    some species adversely affected by a level of 0.1 mg/litre and others
    unaffected by 100 mg/litre; effects on different species do not vary
    consistently with the degree of chlorination of the mixtures. Almost
    all of the studies of the effects of PCBs on aquatic organisms have
    been concerned with Aroclor mixtures. Results have been extremely
    variable with no consistent relationship between percentage
    chlorination or environmental conditions and toxicity, even with
    closely-related organisms. Over 96 h, under static conditions, LC50
    values have ranged between 12 g/litre and >10 mg/litre for various
    aquatic invertebrate species and different Aroclor mixtures.

    Flow-through conditions increased the toxicity of the PCBs. Generally,
    the most toxic mixtures were Aroclors in the mid-range of
    chlorination; low and high percentage chlorination mixtures were less
    toxic. This was also true for sub-lethal effects, such as reproduction
    effects in  Daphnia. Crustaceans seem to be more susceptible to PCBs
    during moult. In model populations, the community structure of
    estuarine species changed on exposure to Aroclor 1254, with the
    numbers of amphipods, bryozoans, crabs, and molluscs decreasing and
    those of annelids, brachyopods, coelenterates, echinoderms, and
    nemerines unaffected. Too few of the groups have been included in
    acute tests to determine whether the results represent variation in
    susceptibility to PCBs or differences in interaction between species.

    There is a similar variation in the toxicity of PCB mixtures for fish,
    with 96-h LC50s varying between 0.008 and >100 mg/litre. Long-term
    tests have shown that acute exposure, particularly in static
    conditions, considerably underestimates the toxicity of the PCB.
    Rainbow trout was particularly susceptible, with embryo-larval stages
    showing a 22-day LC50 of 0.32 g/litre for Aroclor 1254 and a
    no-observed-effect level (NOEL) over 22 days of 0.01 g/litre for
    Aroclors 1016, 1242, and 1254.

    Freshwater fathead minnow showed NOELs of 5.4, 0.1, 1.8, and
    1.3 g/litre for Aroclors 1242, 1248, 1254, and 1260, respectively;
    the estuarine sheephead minnow showed NOELs of 3.4 and 0.06 g/litre
    for Aroclors 1016 and 1254, respectively.

    Experimental evidence has confirmed field observations demonstrating
    reproductive impairment in seals fed on fish containing PCBs
    accumulated in the wild. The effect occurs late in reproduction,
    preventing implantation of the embryo in the uterine wall.

    In short-term tests, the toxicity of Aroclor for birds increased with
    increasing percentage chlorination; 5-day dietary LC50s ranged from
    604 to >6000 mg/kg diet. The main reproductive effects of PCBs on
    birds were reduced hatchability of eggs and embryotoxicity. These
    effects continued after dosing ended, as the hens reduced their PCB
    load via the eggs. There is no evidence that Aroclors cause egg-shell
    thinning, directly; effects on the food consumption and body weight of
    hens have an indirect effect on shell thickness. Sub-lethal effects on
    behaviour and hormone secretion have been reported.

    The acute toxicity of Aroclors for mink decreases with increasing
    percentage chlorination, acute oral LD50s varying between >750 and
    4000 mg/kg body weight; the ferret is less sensitive. Aroclor reduces
    food consumption and, thus, the growth rate of young mink.
    Reproduction of mink is reduced or eliminated by Aroclors, either
    given directly or as natural contaminants in fish. Higher percentage
    chlorinated Aroclors (notably 1254) have a greater effect. The
    reproductive rate returns to normal after cessation of feeding with

    Bats are susceptible to Aroclor released from fat during migration.

    Because the great majority of laboratory tests on aquatic and
    terrestrial organisms were carried out using PCB mixtures, it is not
    possible to identify which specific components of the mixtures were
    responsible for effects. Similarly, because tests were conducted in
    environmentally unrealistic conditions (e.g., beyond the solubility of
    congeners and without sediment present in aquatic tests), it is
    difficult to extrapolate from laboratory to field. However, it can
    reasonably be assumed that any effects on populations of organisms,
    likely to occur more generally in the environment in the future, will
    already have been observed in local populations exposed to high PCB
    levels in the past.  Field studies

    Results suggesting effects of PCBs on fish populations in the field
    are inconclusive. Interpretation of field data on birds is difficult,
    since residues of many different organochlorines are also present.
    Most authors have shown a correlation between effects (embryotoxicity)
    and total organochlorine residues. Of the organochlorine compounds
    present, PCB residues correlate best with the effects on embryos, but
    the results cannot be regarded as proved field effects of the PCBs.

    There is evidence (confirmed in laboratory studies) that PCBs reduce
    the reproductive capacity of sea mammals. The effect is on the
    implantation of the embryo, but there can also be physical changes in
    the female reproductive tract.

    Extrapolation from laboratory, acute and short-term tests to effects
    at the population level in the field is not possible. Uncertainties
    about which components of the PCB mixtures cause effects, the specific
    congeners present in the environment, and the bioavailability of PCB
    components to organisms, all combine to make estimates of likely
    environmental exposures and effects difficult. The effects on sea
    mammal populations can be regarded as proved, but the component(s) of
    the PCB mixtures that are responsible are not yet known.

    Given the trends towards increased contamination of the marine
    environment, attention should be concentrated on the effects on marine
    organisms. There is clear laboratory and field evidence of
    reproductive effects on populations of sea mammals in heavily-polluted
    areas. The residues and effects of PCBs on other populations of sea
    mammals are likely to increase in the future. It is less clear whether
    effects will be seen in other organisms, such as birds feeding on
    marine prey.

    Population and community effects on lower organisms, phytoplankton,
    and zooplankton, would be expected to occur on the basis of laboratory
    experiments. Both the extent and significance of such changes are
    difficult to assess. From currently available information, effects on
    fish populations would not be expected, though fish will act as a
    route of exposure of fish-eating mammals and birds.

    Previously reported effects on terrestrial species, fish-eating,
    freshwater mammals and migratory bats, for example, should be less
    evident as residues of PCBs are redistributed. Residues in terrestrial
    biota currently show little decline overall, but information on
    changes in congeners is scarce or absent. Declines in higher
    chlorinated congeners would be expected to be slow.

    1.1.9  Effects on experimental animals and in vitro systems  Single exposure

    The acute toxicity of Aroclors, after a single oral exposure, is
    generally low in rats. Young animals appear to be more sensitive
    (LD50: 1.3-2.5 g/kg body weight) than adults (LD50: 4-11 g/kg body
    weight). The lowest LD50 reported for Aroclor 1254 in adult rats was
    1.0 g/kg body weight. No differences between the sexes were observed.

    The dermal LD50 in rabbits ranged from >1.26 to <2 g/kg body weight
    for Aroclor 1260 (in corn oil) and from 0.79 to <3.17 g/kg body
    weight for some other undiluted PCB mixtures. With intravenous
    application, an LD50 of 0.4 g/kg body weight for Aroclor 1254 was
    shown in rats; the LD50 after intraperitoneal injection in the mouse
    varied from 0.9 to 1.2 g/kg body weight.  Short-term exposure

    The main targets in mammals, with short-term, oral exposure to PCB
    mixtures or congeners, were the liver, the skin, the immune system,
    and the reproductive system. The Rhesus monkey was the most sensitive
    species tested, females being more sensitive than males. Adult female
    Rhesus monkeys exposed to a diet containing Aroclor 1248 at a level of
    2.5 mg/kg, or 0.09 mg/kg body weight per day, for 6 months, showed an
    increased mortality rate, growth retardation, alopecia, acne, swelling
    of the Meibomian glands, and possibly immunosuppression.

    Microscopically, enlarged fatty liver with focal necrosis, and
    epithelial hyperplasia, and keratinization of hair follicles were
    found. At higher exposure levels, microscopic changes have also been
    observed in other epithelial tissues, such as the sebaceous and
    Meibomian glands, the gastric mucosa, gall bladder, bile duct, nail
    beds, and the ameloblast. Serum levels of total lipid triglycerides
    and cholesterol were decreased. Short-term exposure to commercial PCB
    mixtures induced an increase in the concentrations of total lipids,
    triglycerides, cholesterol, and/or phospholipids in the liver. Among
    the PCB congeners, 3,4,3',4'-tetrachlorobiphenyl 3,4,5,3',4',5'-, and
    2,4,6,2',4',6'-hexachlorobiphenyl were the most potent. Aroclor 1254,
    at a dose level of 0.2 mg/kg body weight per day, also showed several
    other effects, such as lymphoreticular lesions, fingernail detachment,
    and gingival effects, but no acne and alopecia. A NOEL for the general
    toxicity of Aroclor 1242 of 0.04 mg/kg body weight per day was
    established in Rhesus monkeys. Relatively mild effects were shown in
    suckling Rhesus monkeys, exposed to a much higher dose of Aroclor 1248
    of 35 mg/kg body weight per day. Effects in the liver have been best
    investigated in rats and include hypertrophy, fatty degeneration,
    proliferation of the endoplasmic reticulum, porphyria, adenofibrosis,
    bile-duct hyperplasia, cysts, and preneoplastic and neoplastic
    changes. In studies on rats and mice, individual PCB congeners caused
    effects in the liver, spleen, and thymus, the planar congeners being
    most toxic. In monkeys, planar congeners, at doses of 1-3 mg/kg diet,
    induced effects similar in character and severity to those produced by
    Aroclor 1242, at a dose of 100 mg/kg diet, and Aroclor 1248, at a dose
    of 25 mg/kg diet.

    Following dermal exposure of rabbits and mice, PCB mixtures and some
    congeners caused effects on the skin and liver, similar to those found
    after oral exposure. In rabbits, thymic atrophy, a reduction of
    germinal centres of the lymph nodes, and leukopenia were also

    1.1.10  Reproduction, embryotoxicity, and teratogenicity  Reproduction and embryotoxicity

    Comprehensive reproduction and teratogenicity studies have not been
    conducted. In a 2-generation reproduction study on rats, a NOEL of
    0.32 mg/kg body weight, based on reproductive parameters (Aroclor
    1254) and a NOEL of 7.5 mg/kg body weight (Aroclor 1260) were
    established. However, the lowest tested dose of 0.06 mg/kg body weight
    resulted in increased relative liver weights in weanlings.

    In Rhesus monkeys exposed to Aroclor 1016, a NOEL of 0.03 mg/kg body
    weight was established, on the basis of reproductive parameters.
    However, at this level, decreased birth weight was observed and the
    lowest dose tested, of 0.01 mg/kg body weight, resulted in skin

    For Aroclor 1248 (containing PCDFs), a NOEL of 0.09 mg/kg body weight
    was established in Rhesus monkeys, 1 year after exposure ceased.  Teratogenicity

    Available studies on rats and monkeys did not indicate any teratogenic
    effects, when animals were dosed orally during organogenesis. A NOEL
    of 50 mg/kg body weight for Aroclor 1254 was demonstrated in rats with
    regard to pup weight, and a LOEL of 2.5 mg/kg body weight, on the
    basis of fetotoxicity (lesion in thyroid follicular cells) could be

    In teratogenicity tests with individual congeners on mice, rats, and
    Rhesus monkeys, no NOEL was demonstrated. In Rhesus monkeys a dose of
    0.07 mg/kg body weight resulted in maternal toxic effects

    1.1.11  Mutagenicity

    PCB mixtures did not cause mutation or chromosomal damage in a variety
    of test systems. Chromosome breakage was induced in human lymphocytes
     in vitro by 3,4,3',4'-tetrachlorobiphenyl. High concentrations of
    PCB mixtures may cause primary DNA damage, as evidenced by DNA single
    strand breaks in alkaline elution assays.

    1.1.12  Carcinogenicity

    The interpretation of the available animal data involving commercial
    PCB mixtures is often complicated by lack of information concerning
    the presence, or contribution, of chlorinated dibenzofuran impurities
    as well as variations in congener composition.

    A number of long-term carcinogenicity studies have been carried out on
    mice and rats. The PCB mixtures used were Kanechlors 300, 400, and
    500, Aroclors 1254 and 1260, and Clophens A30 and A60. The Clophens
    were reported to be free of PCDFs, but no data were provided on the
    purity of the other PCB mixtures.

    A significant increase in hepatocellular adenomas and/or carcinomas
    was observed in mice fed a diet containing Kanechlor 500 and Aroclor
    1254 at dose levels of approximately 15-25 mg/kg body weight. No
    neoplasms could be detected in mice treated with Kanechlors 300 and

    In rats, an increase in hepatocellular adenomas and/or carcinomas was
    noted in studies on Aroclors 1254 and 1260, and Clophen A30, with an
    exposure period of more than one year. The increase in the incidence
    of tumour-bearing animals in these studies was not considered to be
    statistically significant, however, it was in the case of 2 other
    studies. An increase in the incidence of hepatocellular (trabecular)
    carcinomas and adenocarcinomas was demonstrated with Aroclor 1260 and
    Clophen A60 administered at a dose level of approximately 5 mg/kg body

    The liver tumours concerned were considered to be non-aggressive
    (benign or of low malignancy, no metastasis) and not life shortening.
    Adenofibrosis, a preneoplastic lesion and/or neoplastic nodules in the
    liver were reported in some of the studies. In one test with Aroclor
    1254, a dose-related increase in intestinal metaplasia and
    adenocarcinomas of the glandular stomach was demonstrated in the rat.

    There is a substantial body of evidence to support the enhancing
    effects of PCBs on liver carcinogenesis in rodents pretreated with
    hepatocarcinogens. There is weak evidence for the initiating activity
    of PCB-mixtures in rodents. From the genotoxicity studies reported, it
    can be concluded that PCB-mixtures can be regarded as non-genotoxic.
    These results imply that the association of liver tumours with the
    administration of PCBs in rodents is attributable to some epigenetic
    mechanisms involving enforcement of cell proliferation in the liver
    and other manifestations of liver toxicity, hence a threshold approach
    can be followed in the evaluation of PCB toxicity. The possibility
    that PCBs might enhance carcinogenesis in tissues other than the
    liver, in animals pre-exposed to various tissue-specific carcinogens,
    needs to be addressed. The anticarcinogenic activity of PCBs shown in
    some studies, where PCBs were given to animals during, and prior to,
    the administration of carcinogens, may be related to the microsomal,
    enzyme-inducing properties of PCBs resulting in an increase in

    Overall, there is reason to exercise caution in extrapolating the
    available animal data on the carcinogenic potential of PCBs to humans.

    1.1.13  Special studies

    Lesions induced after exposure to PCB mixtures or individual congeners
    concern the liver, skin, immune system, reproductive system, oedema
    and disturbances of the gastrointestinal tract, and thyroid gland.

    PCBs are able to induce various enzymes in the liver. This has been
    demonstrated, in rats, mice, guinea-pigs, rabbits, dogs, and monkeys,
    for Aroclors 1248, 1254, 1260, and Kanechlor 400 (induction of
    cytochrome P450 and P448). The inducing ability increases with the
    chlorine content in the molecule. It is also dependent on the congener
    composition, congeners with chlorine in the  para- and  meta-
    position inducing the P450 enzyme. For AHH induction, the position of
    the chlorine seems to be more important than the degree of
    chlorination. Congeners with both  para- and at least two  meta-
    positions substituted by chlorine, are the most potent inducers of
    AHH. Distinct inter-species variations have been demonstrated. The
    lowest NOEL (0.025 mg/kg body weight) was found for Aroclor 1260 in
    Osborn-Mendel rats.

    Effects on the endocrine system are seen as alterations in hormonal
    receptor binding and in steroid hormone balance. Direct and indirect
    evidence for a weak estrogenic activity was observed for various
    Aroclors. Decreased levels of gonadal hormones and increased relative
    testes weight were found in rats exposed to 75 mg Aroclor 1242/kg diet
    for 36 weeks. Decreased plasma corticosteroid levels without increased
    adrenal weight, was found in female mice exposed to Aroclor 1254
    (25 mg/kg diet) for 3 weeks. Increased adrenal weight was found in
    another strain given a diet containing 200 mg/kg for 2 weeks.

    PCB mixtures have shown an immunosuppressive effect in various animal
    species, monkeys and rabbits being the most sensitive. The lowest NOEL
    in monkeys was 0.1 mg/kg body weight, and, in rabbits, 0.18 mg/kg body

    Depressed motor-activity was seen in mice administered a single oral
    dose of 500 mg Aroclor 1254/kg body weight. This was probably in
    relation to inhibition of the uptake and release of neurotransmitters.

    PCB mixtures were found to decrease the levels of vitamins A and B1
    in the blood and liver of rats. Decreased levels of vitamins A, B1,
    B2, and B6 were seen in rats and mice exposed to PCB mixtures.

    1.1.14  Factors modifying toxicity, mode of action

    Commercial PCBs show a spectrum of toxic responses, partly resembling
    that of PCDDs and PCDFs. In addition, the analogous structure-activity
    relations of PCB congeners, with respect to most of their toxic
    responses and to their potency in inducing P448-dependent AHH,
    indicate that PCB congeners that are approximate stereoisomers of
    2,3,7,8,-TCDD are the most active. These findings suggest a common
    mechanism of action based on the affinity of these compounds for the
    cytosolic AH-receptor protein. Toxic equivalence factors relating to
    2,3,7,8-TCDD have been proposed for these coplanar PCB congeners. The
    nature of the likely interactions between PCBs, PCDFs, and PCDDs has
    not been adequately investigated. As PCBs can stimulate microsomal
    enzyme activity, they can influence the action of other chemicals that
    undergo microsomal metabolism. Other so-called, non-planar PCB
    congeners may cause other more subtle toxicities. In addition, PCB
    congeners, especially the lower chlorinated ones, may be metabolized
    through arene oxide intermediates and methylsulfonyl metabolites.

    1.1.15  Effects on humans

    The toxicological evaluation of PCBs presents many problems. PCBs
    usually occur as mixtures of many congeners, and many of the data on
    the toxicity of the PCBs are based on the testing of these mixtures.
    Some components of the mixtures are more easily degraded in the
    environment than others. Thus, the general population may be exposed
    to mixtures that are different from those to which workers, working
    with PCBs, are exposed.

    The general population is exposed to PCBs mainly through contaminated
    food (aquatic organisms, meat and dairy products). The daily intake of
    PCBs is of the order of some micrograms per person for most of the
    industrialized countries. Such exposures have not been associated with
    disease. The infant is exposed to PCBs through its mother's milk.
    Daily intake of PCBs may be some micrograms/kg body weight.

    There are great difficulties in assessing human health effects
    separately for PCBs, PCDFs, or PCDDs, since, quite frequently, PCB
    mixtures contain PCDFs. The presence of PCDDs has also been seen
    occasionally, in accidents with certain mixtures. Commercial PCBs have
    been shown to be contaminated with PCDFs and, therefore, in many
    cases, it is not clear which effects are attributable to the PCBs
    themselves and which to the much more toxic PCDFs. Thus, much of the
    data that can be retrieved from large episodes of intoxication in
    humans, e.g., the Yusho, Yu-Cheng, and other intoxications, probably
    reflect effects of exposure to both PCDFs and PCBs.

    The signs of intoxication in Yusho and Yu-Cheng patients were
    hypersecretion of the Meibomian glands of the eyes, swelling of the
    eyelids and pigmentation of the nails and mucous membranes,
    occasionally associated with fatigue, nausea, and vomiting. This was
    usually followed by hyperkeratosis and darkening of the skin with
    follicular enlargement and acneiform eruptions. Furthermore, oedema of
    the arms and legs, liver enlargement and liver disorders, central
    nervous disturbances, respiratory problems e.g., bronchitis-like
    disturbances, and changes in the immune status of the patients were
    also observed. In children of Yusho- and Yu-Cheng patients, diminished
    growth, dark pigmentation of the skin and mucous membranes, gingival
    hyperplasia, xenophthalmic oedematous eyes, dentition at birth,
    abnormal calcification of the skull, rocker bottom heel, and a high
    incidence of low birth weight were observed. Whether or not a
    correlation existed between the exposure and the occurrence of
    malignant neoplasms in these patients could not be definitely
    concluded, because the number of deaths was too small. However, a
    statistically significant increase was observed in male patients, with
    regard to mortality from all neoplasms, liver and lung cancer.

    Under occupational conditions, skin rashes occurred a few hours after
    acute exposure. Furthermore, itching, burning sensations, irritation
    of the conjunctivae, pigmentation the fingers and nails, and chloracne
    were found after exposure to high PCB concentrations. Chloracne is one
    of the most prevalent findings among PCB-exposed workers. Besides
    these dermal signs of intoxication, different authors have found liver
    disturbances, immunosuppressive changes, transient irritation of the
    mucous membranes of the respiratory tract, neurological and unspecific
    psychological or psychosomatic effects, such as headache, dizziness,
    depression, sleep and memory disturbances, nervousness, fatigue, and
    impotence. The overall conclusion is that continuous occupational
    exposure to high PCB and PCDF concentrations may result in effects on
    the skin and liver.

    Two large mortality studies were carried out on cohorts of workers.
    When exposure to Aroclor 1254, 1242, and 1016 occurred, increased
    mortality from cancer of the liver and gall bladder was observed in
    one study and from neoplasms and cancer of the gastrointestinal tract
    in the other. None of the available epidemiological studies provide
    conclusive evidence of an association between PCB exposure and
    increased cancer mortality, because of the small number of deaths in
    exposed populations, the lack of dose-response relationships, and the
    problem of contaminants in the PCB mixtures.

    1.2   Conclusions

    1.2.1  Distribution

    Because of their physical and chemical properties, PCBs have become
    dispersed globally, throughout the environment.

    PCBs are almost universally present in organisms in the environment
    and are readily bioaccumulated. Biomagnification in food chains has
    also been demonstrated.

    Higher chlorinated congeners accumulate preferentially.

    1.2.2  Effects on experimental animals

    The results of animal studies suggest that PCBs are immunosuppressive,
    as assessed by alterations in gross measures of immune function
    (spleen weight, thymus weight, and lymphocyte counts). NOELs in
    monkeys have been estimated at 100 g/kg for Aroclor 1248 and
    <100 g/kg body weight for Aroclor 1254. Immunosuppression appears to
    be a congener-specific effect.

    Reproductive toxicity is, in general, only seen at doses producing
    systemic toxicity in the mother. Neonates feeding on contaminated
    mother's milk (in monkeys and other animal species, used as models)
    appear to be particularly sensitive to PCBs and show reduced growth
    with other toxic symptoms. The NOEL for Aroclor 1016 on reproductive
    effects is 30 g/kg body weight for monkeys; no NOEL could be
    established for the reproductive effects of Aroclor 1248.

    PCBs are not genotoxic and there is inconclusive evidence for action
    as tumour initiators. PCBs do act as tumour promoters. It can be
    concluded that the toxicity of PCB mixtures can be evaluated on a
    threshold basis.

    1.2.3  Effects on humans

    Exposure of the general population to PCBs will be principally through
    food items. Babies will be exposed through the mother's milk.

    Two large episodes of intoxication in humans have occurred in Japan
    (Yusho) and Province of Taiwan (Yu-Cheng). The main symptoms in Yusho
    and Yu-Cheng patients have frequently been attributed to contaminants
    in the PCB mixtures, specifically, to PCDFs. The Task Group concluded
    that symptoms may have been caused by the combined exposure to PCBs
    and PCDFs. However, some of the symptoms, principally, the chronic
    respiratory effects, may have been caused specifically by the
    methylsulfone metabolites of certain PCB congeners.

    1.2.4  Effects on the environment

    While there have been reports of effects on local populations of
    birds, the most important effect of PCBs on organisms in the
    environment has been reproductive failure in sea mammals. This has
    been observed principally in semi-enclosed seas and has led to
    population declines, locally. The prediction that residues of PCBs in
    the environment will gradually be redistributed towards the marine
    environment indicates an increasing hazard for sea mammals in the

    1.3  Recommendations

    *   International agreement on analytical procedures to improve the
        comparability of results of monitoring programmes is recommended.
        Methodology for congener-specific analysis should continue to be
        developed, though the value of analysis based on mixtures is

    *   In order to ensure the reliability of analytical data,
        inter-laboratory quality control studies are strongly recommended.
        It is also recommended that an international network of technical
        support and supervision is established, to allow developing
        countries to participate in monitoring.

    *   Long-term studies using specific congeners, and studies on the
        mechanism of action of constituents of PCBs mixtures, with special
        regard to tumour promotion, are recommended to improve the
        precision of the risk assessment of PCBs.

    *   Epidemiological studies to better assess the risk to neonates are
        required, since new-born infants appear to be the most vulnerable
        sector of the general population, because of high exposure through

    *   Sensitive and specific biomarkers for some of the more subtle
        types of PCB toxicity (such as reproductive, immunological, and
        neural toxicity) should be developed for use in future
        epidemiological studies.

    *   Disposal of PCBs should be carried out by incineration in properly
        designed and run facilities that can guarantee the constant high
        temperatures (above 1000C), residence time, and turbulence
        necessary to ensure complete breakdown.

    *   Methods to remove PCBs already contained in landfills should be

    *   Monitoring of PCBs in the environment and in wildlife should be
        encouraged globally, to follow the expected redistribution of
        residues already present.

    *   Marine mammals are susceptible to reproductive failure as a result
        of PCB contamination. Studies on the population size and
        reproductive success of cetaceans should be encouraged, together
        with further research to establish which congeners are responsible
        for the effects.


    2.1  Identity

    2.1.1  Chemical formula and structure

    The chlorination of biphenyl can lead to the replacement of 1-10
    hydrogen atoms by chlorine; the conventional numbering of substituent
    positions is shown in the diagram:


    The chemical formula can be presented as C12H10-nCln, where n, the
    number of chlorine atoms in the molecule, can range from 1 to 10.

    2.1.2  Relative molecular mass

    The relative molecular mass depends on the degree of substitution.

    Monochlorobiphenyl has a relative molecular mass of 188, while
    completely chlorinated biphenyl (C12Cl10) has a relative molecular
    mass of 494 (US EPA, 1980).

    2.1.3  Common name

    Common name:                polychlorinated biphenyls (PCBs)
    CAS Registry number:        1336-36-3
    RTECS Registry number:      TQ 1350000

    2.1.4  Chemical composition

    The PCBs are chlorinated hydrocarbons, manufactured commercially by
    the progressive chlorination of biphenyl in the presence of a suitable
    catalyst (e.g., iron chloride). Depending on the reaction conditions,
    the degree of chlorination can vary between 21 and 68% (w/w). The
    yield is always a mixture of different isomers and congeners. Thus, a
    total of 209 theoretically different chemical components exist, but
    only about 130 of these are likely to occur in commercial products or
    mixtures of such compounds (Safe, 1990).

    Seventy-eight out of the possible 209 PCB congeners can exist as
    rotational isomers that are enantiomeric to each other. Nineteen PCBs,
    of which 9 are components of commercial PCB formulations, have been
    predicted to be stable at room temperature (Kaiser, 1974).

    Puttmann et al. (1988) separated the atropisomers of
    2,3,4,6,2',4'-hexachlorobiphenyl and demonstrated that they possess
    different biological effects with regard to  in vivo enzyme induction
    (aminopyrine  N-demethylase, aldrin epoxidase, cytochrome P-450
    content, morphine UDP-glucuronosyl transferase) in Sprague-Dawley

    Unlike the dioxins or dibenzofurans, the phenyl rings of a PCB are not
    constrained through ring fusions and have relatively unconstrained
    rotational freedom. Chlorines at the  ortho (2,2', 6,6') positions
    introduce constraints on rotational freedom that can hinder
    coplanarity of the phenyl rings. X-ray crystallographic studies
    (McKinney & Singh, 1981) indicate that the preferred conformation for
    all PCBs, including those without  ortho-substituents, is
    noncoplanar. The proportion of molecules of a particular congener
    assuming a coplanar configuration becomes increasingly small as the
    degree of  ortho-substitution and the energetic cost of conforming
    increases. However, PCBs without  ortho-substitution are often
    referred to in the biological literature as the planar (or coplanar)
    PCBs and all others as the nonplanar (or noncoplanar) PCBs. This
    terminology, though somewhat misleading, is also used throughout this
    publication for convenience and ease of referring back to the
    published literature. It is widely recognized that certain biological
    activities of the PCBs vary, at least quantitatively, with
    stereochemical differences in the congeners.

    Individual manufacturers have their own system of identification for
    their products. In the Aroclor series, a 4-digit code is used;
    biphenyls are generally indicated by 12 in the first 2 positions,
    while the last 2 numbers indicate the percentage by weight of chlorine
    in the mixture; thus, Aroclor 1260 is a polychlorinated-biphenyl
    mixture containing 60% of chlorine. An exception to this
    generalization is Aroclor 1016, which is a distillation product of

    Aroclor 1242 containing only 1% of components with 5 or more chlorine
    atoms (Burse et al., 1974). With other commercial products, the codes
    may indicate the approximate mean number of chlorine atoms in the
    components; thus Clophen A60, Phenochlor DP6, and Kanechlor 600 are
    biphenyls with an average of about 6 chlorine atoms per molecule
    (equivalent to 59% chlorine by weight).

    Ballschmiter & Zell (1980) proposed a numbering system for the PCB
    congeners, that was later adopted by the International Union of Pure
    and Applied Chemists (IUPAC). The number, structure, and isomer group
    are given for each congener in the paper of McFarland & Clarke (1989)
    (see Appendix A). In the literature, the structure of a congener is
    given in 2 ways; for example 2,2',5,5' or 2,5,2',5' (No 52).

    Individual PCBs have been synthesized for use as reference samples in
    the identification of gas-liquid chromatographic peaks, for
    toxicological investigations, and for studying their metabolic fate in
    living organisms, for which purpose they have been prepared labelled
    with carbon-14 (Hutzinger et al., 1971; Jensen & Sundstrm, 1974a;
    Sundstrm & Wachtmeister, 1975).

    The proportions of PCBs with 1-9 chlorine substituents in the Aroclors
    are shown in Table 1.

    It is apparent, from gas chromatographic analyses of commercial
    products, that PCB mixtures differ with respect to the individual
    congeners present and their relative concentrations (Jensen &
    Sundstrm, 1974a; Albro & Parker, 1979; Ballschmiter & Zell, 1980;
    Albro et al., 1981; Mullin et al., 1984; Safe et al., 1985a;
    Alford-Stevens, 1986).

    There have been several investigations to identify individual PCBs in
    commercial products. The components of the Aroclors were separated by
    column and gas-liquid chromatography and many of the peaks
    characterized by high-resolution mass spectrometry and nuclear
    magnetic resonance, and also by comparison with synthesized PCBs
    (Table 2) (see also DFG, 1988).

    Jensen & Sundstrm (1974a) recognized that conventional gas-liquid
    chromatography was not suitable for separating all the components, so
    they devised a preliminary fractionation on a charcoal column, which
    separated the component PCBs according to the number of chlorines in
    the 2,6,2' or 6' positions in the molecule ( o-chlorines). They
    compared the gas-liquid chromatographic peaks with those of 90
    synthesized PCBs, and were able to characterize and quantify 60
    components of Clophens A50 and A60.

        Table 1.  Approximate percentages (w/v) of Aroclors with different degrees of

    Number of   Chlorine
    chlorine    weight                              Aroclor
    atoms in    (%)                                                                          
    molecule                 1221    1232    1016   1242    1248   1254    1260

    0             0           10      -       -
    1            18.8         50      26       2      3
    2            31.8         35      29      19     13       2
    3            41.3          4      24      57     28      18
    4            48.6          1      15      22     30      40     11
    5            54.4                                22      36     49      12
    6            59.0                                 4       4     34      38
    7            62.8                                                6      41
    8            66.0                                                        8
    9            68.8                                                        1

    a  From: WHO/EURO (1987).

    2.1.5  Technical product

    Major trade names

    The PCBs manufactured commercially are known by a variety of trade
    names including: Aroclor, Pyranol, Pyroclor (USA), Phenoclor, Pyralene
    (France), Clophen, Elaol (Germany), Kanechlor, Santotherm (Japan),
    Fenchlor, Apirolio (Italy), and Sovol (USSR). Table 3 contains the
    most common trade names for commercial products, some of which are not
    in use any more (Brinkman & De Kok, 1980; WHO/EURO, 1987).

    2.1.6  Purity and impurities

    Commercial PCBs are not sold according to a composition specification,
    but according to their physical properties. The composition of
    Aroclors and Clophens has been presented in recent papers; the
    composition of 5 Aroclors is shown in Tables 1 and 2. In Table 1, the
    approximate composition is expressed as the percentage of chlorine
    weight, and, in Table 2, the composition of the chlorine substitution
    pattern is expressed in mol % (Albro & Parker, 1979; Albro et al.,
    1981; Jones, 1988). The composition of the chlorine substitution
    pattern for 4 Clophens is described by Duinker & Hillebrand (1983) and

    Jones (1988). It should be kept in mind that nothing can be said about
    the variations in the different lots of these mixtures. Impurities
    known to be present in commercial PCBs are chlorinated dibenzofurans
    and chlorinated naphthalenes (Vos et al., 1970; Bowes et al., 1975;
    Albro & Parker, 1979; Albro et al., 1981; Duinker & Hillebrand, 1983;
    Rappe et al., 1985a). The concentrations of PCDFs in Aroclor, Clophen,
    Phenoclor, and Kanechlor are summarized in Tables 4 and 5.

    Different authors have examined the presence of PCDFs in PCB mixtures.
    Bowes et al. (1975) found 0.8-2.0 mg/kg in samples of Aroclor 1248 and
    1260, but none in Aroclor 1016, 8.4 mg/kg in Clophen A60, and
    13.6 mg/kg in Phenoclor DP-6. Rappe et al. (1985a) and Bentley (1983)
    found levels of PCDFs up to 40 mg/kg in a number of commercial PCBs.
    Recently, Wakimoto et al. (1988) found a number of extremely toxic
    PCDFs in several Japanese and American commercial PCB preparations.
    These isomer-specific analyses revealed the 2,3,7,8-tetra-,
    1,2,4,7,8-penta-, 1,2,3,7,8-penta-, 2,3,4,7,8-penta-, and
    1,2,3,6,7,8-hexachlorodibenzofurans. The concentrations in unused
    Kanechlor 300, 400, 500, and 600, were 7.5, 26, 7.2, and 5.4 mg/kg,
    respectively, and those in Aroclors 1242, 1248, 1254, and 1260, were
    0.6, 3.7, 4.2, and 7.5 mg/kg, respectively. Brown et al. (1988) found
    that the electrical use of PCB dielectric fluids in transformers and
    capacitors did not increase the PCDFs content significantly.

    More data about the occurrence of PCDFs in the different commercial
    PCB mixtures are summarized in WHO/EURO (1987).

    There are no reports on the presence of PCDDs in commercial mixtures
    (Bowes et al., 1975). Wakimoto et al. (1988) could not find PCDDs in
    the above samples of Kanechlors and Aroclors with a detection limit of
    <2 g/kg.

    2.2  Physical and chemical properties

    Individual pure PCB congeners are colourless, often crystalline
    compounds, but commercial PCBs are mixtures of these congeners with a
    clear, light yellow or dark colour. They do not crystallize at low
    temperatures, but turn into solid resins. Because of the chlorine
    atoms in the molecule, their density is rather high. PCBs are, in
    practice, fire resistant with rather high flash-points (170-380C).
    They form vapours heavier than air, but do not form any explosive
    mixtures with air. They possess very low electrical conductivity and
    an extremely high resistance to thermal breakdown, and it is on the
    basis of these properties that they are used as cooling liquids in
    electrical equipment (US EPA, 1980; WHO/EURO, 1987; DFG, 1988).

        Table 2.  PCB compositions of aroclors in mol %a

    IUPAC        Chlorine                                        Aroclor
    No.          substitution
                 pattern                 1242        1016        1248        1254      1260

                 BP                      0.01        0.50
    1            2                       0.68        0.80
    2            3                       0.04        0.10
    3            4                       0.22        1.00
    4            2.2'                    3.99        4.36        0.25
    6            2.3'                    1.24        1.37        0.69        0.07
    7            2.4                     1.04        1.16
    8            2.4'                    8.97       10.30        0.18
    9            2.5                     0.31        0.34        trace
    10           2.6                     0.13        0.20
    12           3.4                     0.09        0.11
    13           3.4'                    0.12        0.12
    14           3.5                     0.35        0.37
    15           4.4'                    0.99        1.07
    16           2.3.2'                  3.25        3.50        0.84
    17           2.4.2'                  2.92        3.14        0.19
    18           2.5.2'                  9.36       10.87        9.95        0.07
    19           2.6.2'                  0.97        1.08
    20           2.3.3'                  3.64        3.99
    22           2.3.4'                  2.64        2.80        1.24        trace     trace
    25           2.4.3'                  1.68        1.79
    26           2.5.3'                  0.55        0.62        0.75
    27           2.6.3'                  0.54        0.58
    28           2.4.4'                  13.30      14.48        trace
    31           2.5.4'                  4.53        4.72        9.31        0.72
    32           2.6.4'                  2.15        2.31        1.46
    33           3.4.2'                  2.83        3.08
    35           3.4.3'                  0.66        0.38
    37           3.4.4'                  1.62        1.89        1.28        0.20      0.09
    39           3.5.4'                  1.03        1.08
    40           2.3.2'.3'               0.15        0.18        1.12        0.26      0.04
    41 '                1.67        2.00
    42           2.3.2'.4'                                       7.05        2.18      0.66
    43 '                0.44        0.47
    44           2.3.2'.5'               1.06        1.14
    45 '                0.90        1.00        5.73        0.15
    46           2.3.2'.6'               0.31        0.33
    47           2.4.2'.4'               1.65        1.8         3.18        0.52      0.88
    48 '                1.33        1.41

    Table 2. (cont'd).

    IUPAC        Chlorine                                        Aroclor
    No.          substitution
                 pattern                 1242        1016        1248        1254      1260

    ?            2.5.2'.4'               -           -           3.81        1.63      0.44
    49           2.4.2'.5'               3.28        3.48
    52           2.5.2'.5'               4.08        4.35        8.36        4.36      1.91
    53           2.5.2'.6'               0.97        1.07        6.30        0.13
    54           2.6.2'.6'               0.17        0.19
    55 '                                        0.11        0.43      0.12
    56           2.3.3'.4'               0.60        trace       0.18        0.03
    60 '                0.21
    66           2.4.3'.4'               0.81        0.14        4.95        2.24      0.22
    70           2.5.3'.4'               1.11                    6.38        4.75      0.85
    71           2.6.3'.4'                                       0.65
    72           2.5.3'.5'               0.33                    2.10        1.01      0.28
    74 '                2.02        1.35        0.25        0.30      0.09
    75 '                2.18        2.40
    76 '                trace                   trace       0.18      0.01
    77           3.4.3'.4'               0.34                    0.47        0.12      0.04
    78 '                0.52
    79           3.4.3'.5'               0.24                    trace       0.23      0.04
    80           3.5.3'.5'                                       trace       trace     trace
    81 '                0.28
    83 '.3'                                     trace       0.32      0.09
    84 '.3'             0.38        0.01        0.71        1.72      0.69
    85 '.4'             0.40                    0.55        2.15      0.31
    ?  '.5'                                     0.02        0.55      0.14
    87 '.5'             0.09                    1.05        3.81      1.10
    91 '.4'             trace                   1.78        5.00      3.22
    92 '.5'             0.12                    0.20        0.63      0.21
    95 '.5'             0.53        0.18
    97 '.3'                                     0.78        2.59      0.63
    98 '.3'             0.13        0.04
    99 '.4'             0.55                    2.52        6.10      0.82
    101'.5'             0.27                    1.50        6.98      5.04
    102'.6'                                     trace       trace     trace
    105'.4'             0.25
    106'                                                  0.40      0.06
    108'.5'             0.46        0.16
    110'.4'                                     1.69        8.51      3.57
    113'.5'             0.39        0.01        3.10        trace     0.01
    114'                                                  0.25      0.03
    118'.4'                                                 8.09      2.00
    120'.5'             0.31                    trace       0.15      3.01
    121'.5'             0.92                    4.32        3.51      0.57

    Table 2. (cont'd).

    IUPAC        Chlorine                                        Aroclor
    No.          substitution
                 pattern                 1242        1016        1248        1254      1260

    123'.4'             0.36
    ?  '.3'                                     trace       0.76      1.88
    126'.4'             0.03                                0.16      1.59
    127'.5'             0.05
    128'.3'.4'                                              1.31      0.47
    131'.3'                                               0.14      0.01
    132'.3'.6'                                  trace       2.00      2.77
    133'.3'.5'                                  1.13        0.03      0.06
    134'.3'                                   0.11        0.38      1.01
    135'.3'.6'                                              0.20      0.29
    136'.3'.6'                                  0.20        0.34      1.12
    138'.4'.5'          0.08                    0.19        4.17      5.01
    143'.6'           0.07
    148'.4'.6'                                  0.12        0.07      0.06
    149'.3'.6'                                  0.77        3.59      9.52
    151'.5'                                   trace       0.33      0.06
    153'.4'.5'          0.02                    0.13        3.32      8.22
    154'.6'                                                 0.14
    156'.4'                                                         0.41
    157'.4'.5'                                              0.18      0.03
    158'.4'                                               0.46      0.18
    159'.3'.5'                                              0.75      1.48
    163'.4'                                                         trace
    167'.4'.5'                                              0.21      0.17
    168'.4'.5'                                  0.56        4.23      0.59
    170'.3'.4'                                            0.43      0.62
    171'.3'.4'                                            0.30      4.31
    174'.3'.6'                                            trace     0.09
    176'.3'.6'                                0.09        trace     0.57
    177'.3'.4'                                                      trace
    179'.3'.6'                                            0.56      0.83
    180'.4'.5'                                            0.76      7.20
    181'.4'                                             0.28      2.72
    182'.4'.6'                                            trace     0.47
    183'.4'.5'                                            1.16      2.58
    185'.5'                                             1.11      5.65
    186'.6'                                 trace       trace     0.37
    187'.4'.5'                                            0.48      1.12
    189'.4'.5'                                                      0.13
    190'.4'                                                       0.02
    192'.5'                                             0.20      0.97

    Table 2. (cont'd).

    IUPAC        Chlorine                                        Aroclor
    No.          substitution
                 pattern                 1242        1016        1248        1254      1260

    193'.4'.5'                                            2.30
    194'.3'.4'.5'                                                   2.21
    195'.3'.4'                                                    trace
    196'.3'.4'.6'                                                   0.79
    197'.3'.4'.6'                                                   0.30
    198'.3'.5'                                          1.00      0.15
    199'.3'.6'                                                    0.38
    200'.3'.5'.6'                                         trace     0.15
    202'.3'.5'.6'                                         trace     0.31
    203'.4'.5'                                                    0.08
    204'.4'.6'                                          trace     0.13
    205'.4'.5'                                                    0.01
    206'.3'.4'.5'                                                 0.51
    207'3'.4'.6'                                                  1.15
    208'.3'.5'.6'                                                 1.64
    ?  '.3'.5'.6'                                                 0.18

    a  From: Albro & Parker (1979); Albro et el. (1981).

    Table 3.  The trade marks of PCB products and mixtures containing PCBsa

    Aceclor (t)              Disconon (c)             PCBs
    Apirolio (t,c)           Dk (t,c)                 Phenoclor (t,c)
    Aroclor (t,c)            Duconol (c)              Polychlorinated biphenyl
    Arubren                  Dykanol (t,c)            Polychlorobiphenyl
    Asbestol (t,c)           EEC-18                   Pydraulc
    Askarel                  Elemex (t,c)             Pyralene (t,c)
    Bakola 131 (t,c)         Eucarel                  Pyranol (t,c)
    Biclor (c)               Fenchlor (t,c)           Pyroclor (t)
    Chlorextol (t)           Hivar (c)                Saf-T-Kuhl (t,c)
    Chlorinated Biphenyl     Hydol (t,c)              Santotherm FRb
    Chlorinated Diphenyl     Inclor                   Santovac 1 and 2
    Chlorinol                Inerteen (t,c)           Siclonyl (c)
    Chlorobiphenyl           Kanechlor (t,c)          Solvol (t,c)
    Clophen (t,c)            Kennechlor               Sovol
    Clorphen (t)             Montar                   Therminol FRb
    Delor                    Nepolin
    Diaclor (t,c)            No-Flamol (t,c)
    Dialor (c)               PCB

    a  From: WHO/EURO (1987).
    b  Previous products (FR-series) used as pressure oil contained PCBs, but current
       products are a different series and do not contain PCBs.
    c  Previous products (A-series) e.g., PYDRAUL A-200 contained PCBs, but current
       commercial products are B, C, or D-series and do not contain any chlorinated

      (t)  Used in transformers.
      (c)  Used in capacitors.

    Table 4.  Concentrations of chlorinated dibenzofuransa in Aroclor, Clophen, and

    PCB                      4-Cl           5-Cl           6-Cl         Total

    Aroclor 1248 (1969)      0.5 (25)       1.2 (60)       0.3 (15)      2.0
    Aroclor 1254 (1969)      0.1 (6)        0.2 (12)       1.4 (82)      1.7
    Aroclor 1254 (1970)      0.2 (13)       0.4 (27)       0.9 (60)      1.5
    Aroclor 1260 (1969)      0.1 (10)       0.4 (40)       0.5 (50)      1.0
    Aroclor 1260 (lot AK3)   0.2 (25)       0.3 (38)       0.3 (38)      0.8
    Aroclor 1016 (1972)      ND             ND             ND
    Clophen A-60             1.4 (17)       5.0 (59)       2.2 (26)      8.4
    Phenoclor DP-6           0.7 (5)        10.0 (74)      2.9 (21)     13.6

    a  Expressed as mg PCB/kg. Values in parentheses represent quantity as percentage
       of total dibenzofurans.
    b  From: Bowes et al. (1975).
       ND = not detected (0.001 mg/kg).

    Table 5.  Concentrations of chlorinated dibenzofurans in Kanechlorsa

    Kanechlor                   Chlorodibenzofurans                  Concentration

                   Di-   Tri-   Tetra-   Penta-   Hexa-    Hepta-     b        c

    300                         +        +                            1       1.5
    400            +     +      +        +                           18      17
    500                  +               +        +        +          4       2.5
    600                         +        +        +        +          5       3

    a  From: Nagayama et al. (1975).
    b  Calculated from peak heights.
    c  Calculated by perchlorination method.

    PCBs have a high degree of chemical stability under normal conditions.
    They are very resistant to a range of different oxidants and other
    chemicals. According to laboratory tests, they stay chemically
    unchanged, even in the presence of oxygen or some active metals at
    high temperatures (up to 170C) and for protracted periods (WHO/EURO,

    PCBs are practically insoluble in water, whereas they dissolve easily
    in hydrocarbons, fats, and other organic compounds and they are
    readily absorbed by fatty tissues (WHO/EURO, 1987).

    Some physical and chemical data for a number of Aroclors are presented
    in Table 6.

    Foreman & Bidleman (1985) estimated the liquid phase vapour pressures,
    at 25C, of 134 PCB congeners found in 5 Aroclor fluids, using a
    capillary gas chromatographic method in conjunction with published
    retention indices of PCBs.

    Burkhard et al. (1985) predicted Henry's Law Constants from the ratio
    of the liquid (or subcooled liquid) vapour pressure and aqueous
    solubility for PCB congeners. The predicted values were in fair
    agreement with experimental values and the error for these constants
    was estimated to be a factor of 5 in the temperature range of 0-40C.
    For the PCB congeners, Henry's Law Constants were independent of the
    relative molecular mass and increased approximately an order of
    magnitude with a 25C increase in temperature.

    Aqueous solubility is considered an essential parameter for predicting
    the fate and transport of organic chemicals in the environment. As
    already stated, some physical and chemical data are given for 6
    Aroclor mixtures in Table 6 (Alford-Stevens, 1986). However, during
    the last 5 years, much more information on aqueous solubility, melting
    points, entropies of melting, Henry's law constants, and vapour
    pressures has become available. This information concerns not only PCB
    mixtures but also individual congeners.

    Opperhuizen et al. (1988) studied the aqueous solubilities of 45
    chlorinated biphenyls and the relationships between activity
    coefficient and chemical structure parameters (total surface area
    (TSA) and total molecular volume (TMV)) of hydrophobic chemicals, to
    understand the nature of hydrophobicity. The aqueous solubilities of
    PCBs showed a linear relationship between logarithms of aqueous
    activity coefficients or TSA and TMV.

        Table 6.  Physical and chemical properties of a number of Aroclorsa

    Substance   Water         Vapour          Density    Appearance            Henry's Law     Refractive index        Boiling point
    Aroclor     solubility    pressure        (g/cm3)                          constant                                (distillation
                (mg/litre)    (torr) 25C     25C                             (atm-m3/mol                             range) (750
                25C                                                           at 25C)b                               torr, C)

    1016        0.42          4.0  10-4      1.33       Clear, mobile oil     2.9  10-4      1.6215-1.6235           325-356
                                                                                               (at 25C)

    1221        0.59c         6.7  10-3      1.15       Clear, mobile oil     3.5  10-3      1.617-1.618 (at 20C)   275-320

    1232        0.45          4.1  10-3      1.24       Clear, mobile oil     unknown         unknown                 290-325

    1242        0.24          4.1  10-3      1.35       Clear, mobile oil     5.2  10-4      1.627-1.629 (at 20C)   325-366

    1248        0.054         4.9  10-4      1.41       Clear, mobile oil     2.8  10-3      unknown                 340-375

    1254        0.021         7.7  10-5      1.50       Light yellow          2.0  10-3      1.6375-1.6415           365-390
                                                         viscous oil                           (at 25C)

    1260        0.0027        4.0  10-5      1.58       Light yellow          4.6  10-3      unknown                 385-420
                                                         sticky resin

    a  From: IARC (1978); WHO/EURO (1987); ATSDR (1989).
    b  These Henry's Law Constants were estimated by dividing the vapour pressure by the water solubility. The first water solubility
       given in this table was used for the calculation. The resulting estimated Henry's law constant is only an average for the
       entire mixture; the individual chlorobiphenyl isomers may vary significantly from the average. Burkhard et al. (1985)
       estimated the following Henry's Law Constants (atm-m3/mol) for various Aroclors at 25C: 1221 (2.28  10-4), 1242 (3.43  10-4),
       1248 (4.4  10-4), 1254 (2.83  10-4), 1260 (4.15  10-4).
    c  At 24C.

    Dickhut et al. (1986) studied the solubilities of 6 higher chlorinated
    biphenyl congeners at different temperatures and found that the
    solubility increased exponentially with temperature in the range of
    0.4-80C. From the temperature dependence of solubility, enthalpies of
    solution were calculated. The same results were found by Doucette &
    Andren (1988), who determined the aqueous solubilities of a few PCBs,
    using a generator-column technique, at temperatures of 4.0, 25.0, and

    The dissolution of extremely hydrophobic chemicals that may be
    associated with a relatively constant endothermic enthalpy of solution
    and an endothermic enthalpy of fusion that is proportional to the
    solute's melting point is discussed by Opperhuizen et al. (1987) and
    Dickhut et al. (1987).

    Dunnivant & Elzerman (1988) estimated the aqueous solubilities and
    Henry's Law Constants (HLC) for 26 selected PCB congeners for the
    evaluation of quantitative structure-property relationships (QSPRs).
    Aqueous solubilities (as solids at 25C, column generation technique),
    determined for the 26 congeners, ranged from 1.08  10-5 to
    9.69  10-10 mol/litre and generally decreased with relative molecular
    mass. HLCs (25C, gas purge technique), determined for 20 congeners,
    ranged from 0.3  10-4 to 8.97  10-4 atm.m3/mol. Measured HLCs were
    not correlated with relative molecular mass, but increased with the
    degree of  ortho-chlorine substitution within a relative molecular
    mass class.

    Vapour pressures calculated from the product of solubility (mol/m3)
    and HLC (atm-m3/mol) data, generally decreased with relative
    molecular mass and increased with increasing degree of
     ortho-chlorine substitution (Dunnivant & Elzerman, 1988; Hawker,
    1989). Westcott et al. (1981) used a semimicro gas saturation method
    to determine the vapour pressures of 3 PCB isomers and 2 Aroclor

    Experimental data were tabulated and the relationships between the
    environmentally relevant physical chemical properties of the PCBs
    critically reviewed by Shui & Mackay (1986). Aqueous solubility,
    vapour pressure, Henry's law constant, and octanol-water partition
    coefficient were discussed and recommended values given for 42 of the
    209 congeners; procedures were suggested for estimating the properties
    of the other congeners.

    2.2.1  Log n-octanol/water partition coefficient

    The environmental fate of PCBs is governed primarily by the
    partitioning process. Partitioning processes that are of particular
    interest with regard to environmental problems include: the octanol/
    water partition coefficient and the aqueous solubility. The octanol/
    water partition coefficient is a measure of the hydrophobicity of a
    substance and, in this respect, it has been used to predict the extent
    of bioconcentration of organic pollutants in organisms. Miller et al.
    (1984) studied the octanol/water partition coefficients for 16 PCBs
    and Hawker & Connell (1988) for 13 PCB congeners, using the generator
    column method. These partition coefficients were used to confirm a
    highly significant linear relationship between log Kow and the
    logarithm of the relative retention time on a nonselective gas
    chromatographic stationary phase. The total surface areas (TSA) for
    all the PCB congeners were determined by assuming planar molecules,
    van der Waal's radii for component atoms, and appropriate values for
    solvent radius, bond angles, and distances. The TSA was highly
    significantly correlated with log Kow and the relationship was used
    to calculate log Kow values for all the PCB congeners. In the report
    of Hawker & Connell (1988), log Kow values are summarized for all 209
    PCB congeners. These log Kow values range from 4.46 to 8.18.

    2.2.2  Conversion factorsa

    1016                                1 mg/m3 = 0.095 ppm
    1221                                1 mg/m3 = 0.12  ppm
    1232                                1 mg/m3 = 0.105 ppm
    1242                                1 mg/m3 = 0.092 ppm
    1248                                1 mg/m3 = 0.008 ppm
    1254                                1 mg/m3 = 0.075 ppm
    1260                                1 mg/m3 = 0.065 ppm

    2.3  Analytical methods

    Reviews have been published on the methods used for the determination
    of organochlorine compounds including PCBs in environmental samples
    (Panel on Hazardous Trace Substances, 1972; Holden, 1973; US DHEW,
    1978; Slorach & Vaz, 1983; Jensen, 1984, 1985; Erickson 1985;
    Alford-Stevens, 1986; NIOSH, 1987; DFG, 1988; WHO/EURO, 1987, 1988).


    a  These air conversion factors were calculated by using the average
       molecular mass at 25C.

    No two laboratories used identical methods, though all the methods
    have features in common. The techniques appear to be those previously
    developed for the determination of organochlorine pesticides, with
    appropriate modifications for the presence of PCBs, and the studies on
    PCBs sometimes form part of a wider programme for monitoring
    persistent organochlorine compounds in the environment. In the past,
    the major difficulty in the determination of PCBs was to obtain a
    single quantitative figure from a variable mixture of components. The
    PCBs were chlorinated with antimony pentachloride to decachloro-
    biphenyl, which was measured as a single peak (Greve & Wegman, 1983;
    Tuinstra, 1983). At the moment, chemists and toxicologists are no
    longer trying to derive a single quantitative figure, preferring
    instead to quantify individual congeners. The legislation in certain
    countries is now based on quantifying a few selected congeners,
    instead of reporting "total PCBs". It is also felt that for
    pinpointing areas with high levels of contamination, in order to rank
    them into low, medium, or high priority areas for action, highly
    accurate laboratory analyses are not necessary; instead, analytical
    competence and the use of adequate controls and standards, resulting
    in consistent, reasonably accurate results would be enough. Of course,
    for complicated research, especially involving laboratories in
    different countries, standardization of techniques through
    collaborative and comparative studies would be necessary.

    Jones (1988) and Safe et al. (1985a) studied the occurrence of
    specific PCB congeners in commercial formulations or mixtures. The
    congener composition of commercial formulations differs from
    batch-to-batch, between manufacturing processes, and with the level of
    chlorination. The presence of congeners in the environment will depend
    on the eventual use of commercial formulations, the quantity of each
    formulation manufactured, as well as on the isomer composition of the

    On the basis of a literature review of the occurrence of PCB congeners
    in environmental and biological samples and human tissues, and
    consideration of the relative toxicity and persistence of the
    congeners, suggestions were made by Jones (1988), with regard to the
    most relevant components to be quantified in human foodstuffs and
    tissues, using a selective analytical approach.

    The congeners reported (Safe et al., 1985a; Duinker et al., 1988;
    McFarland & Clarke, 1989) as being the most abundant in human tissues
    and which are most important, are compounds with IUPAC numbers 28, 52,
    74, 77, 99, 101, 105, 118, 126, 128, 138, 153, 156, 169, 170, 179, and
    180 (comprising >70% of total PCBs and being of greatest

    toxicological significance). Because of their reported occurrence or
    toxicity, congeners with IUPAC numbers 8, 37, 44, 49, 60, 66, 70, 82,
    87, 114, 158, 166, 183, 187, and 189 might also be considered. Duinker
    et al. (1988) were also of the opinion that toxicity should be
    considered as a criterion for the selection of PCB congeners for
    analysis in environmental samples. Most of these congeners can be
    accurately determined with the application of the multidimensional,
    high-resolution GC-ECD techniques.

    PCB reference materials are necessary for the qualitative and
    quantitative calibration of analytical apparatus and methods (e.g.,
    determination of retention times, response factors, and reference
    spectra in chromatographic and spectroscopic analyses) and for the
    study of biological activity. Lindsey & Wagstaffe (1989) described the
    production and certification of 10 high-purity PCBs with IUPAC numbers
    8, 20, 28, 35, 52, 101, 118, 138, 153, and 180.

    Mes et al. (1989a) described an analytical method to determine 34
    isomers of PCB congeners in fatty foods. A sample was extracted with
    an acetone:hexane mixture and the extracts washed and dried; this was
    followed by a clean-up and determination by gas chromatography. GC/MS
    was used for confirmation.

    Environmental PCB residues are often expressed in terms of relative
    Aroclor composition. Schwartz et al. (1987) assessed the similarity of
    Aroclors with class models derived for fish and turtles, to ascertain
    if the PCB residues in the samples could be described by an Aroclor or
    Aroclor mixture. The PCB residues in fish and turtles were analysed
    with Soft Independent Modelling of Class Analogy, a principal
    components analysis (PCA) technique. Using PCA, it was inappropriate
    to report these samples as an Aroclor or Aroclor mixture.

    2.3.1  Sampling strategy and sampling methods

    The quality and usefulness of analytical data, especially in the
    microgram-nanogram range, or even lower, depend critically on the
    validity of the sample and the adequacy of the sampling programme. 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
    representative of the situation of interest (WHO/EURO, 1987). Slorach
    (1984) described the problems encountered with the sampling and
    determination of PCBs in breast milk (see also WHO/EURO, 1985, 1988).

    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,
    transportation, and storage (WHO/EURO, 1987).

    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 (WHO/EURO, 1987).

    Because environmental samples are typically heterogeneous, a
    sufficiently large number of samples (10 or more) must be analysed to
    obtain meaningful composition data. The number of individual samples
    that should be analysed will depend on the kind of information
    required. If an average composition 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 are analysed. If composition profiles,
    time trends, or the variability of the sample population is of
    interest, many samples need to be collected and analysed 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
    analyte. In addition, measurements should be made to ascertain
    whether, and to what extent, any reagent or solvent used may
    contribute or interfere with the analytical results (laboratory and
    solvent blanks). The recovery tests are frequently used and are
    necessary to evaluate the analytical methodology. Uncontaminated
    samples from control sites that have been spiked with the analyte of
    interest provide the best information, because they simulate any
    matrix effect. When feasible, isotopically labelled (13C, 37Cl)
    analytes spiked into the sample provide the greatest accuracy, since
    they are subjected to the same matrix effects as the analytes. The
    13C-labelled compounds can be used to:

     (a) validate sampling (sampling surrogate);

     (b) validate analytical waste (clean-up surrogate);

     (c) validate quantification (internal standard).

    Only a small number of 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 a few
    cases (WHO/EURO, 1987).  Extraction procedures


    The sampling device used to collect and determine PCBs in air consists
    of a glass fibre filter and a Florisil stick. The glass fibre filter,
    held in a stainless steel holder, removes particles larger than
    0.3 m. The air passes from the filter to the Florisil stick, which is
    made in 2 sections, to provide information on migration and trapping
    efficiency for PCBs. Each section contains 0.4 g of Florisil preceded
    and followed by a glass wool plug. The front and back sections are
    separated by 2 plugs of glass wool. The front is spiked with 0.1 g of
    p,p'-DDE as a surrogate for recovery measurement and as an indication
    of analyte migration. The detection limit for PCBs in air is reported
    to be 0.3 ng/m3 (Anon., 1985; WHO/EURO, 1987; NIOSH, 1987).

    Particulate fallout from air has been trapped on 200 m nylon net
    coated with silicone oil, and the PCBs then extracted with hexane
    (Sdergren, 1972). Separate determinations of particulate and vapour
    phase PCBs in air have been made by passing a large volume of air
    through a filter followed by an impinger containing hexane or toluene
    (Rappe et al., 1985c), a polyurethane plug (Bidleman & Olney, 1974),
    or ceramic saddles coated with OV 17 silicone (Harvey & Steinhauer,
    1974) to absorb the vapour.

    Surface sampling

    Surface sampling of PCBs can be carried out using a wet-wipe procedure
    with a cotton gauze pad that has been dampened with hexane before
    collecting the sample. The sampled area is 0.25 m2. The wet-wipe
    sampling procedure collects both the contaminants from the surface and
    the contaminants that can be extracted from pores in the material.
    Materials such as waxes and plasticizers may interfere with the
    chemical analysis (WHO/EURO, 1987).

    Another sampling method has been described by Rappe et al. (1985c),
    where a dry filter paper or Kleenex tissue is used first, for wiping,
    followed by a wet wipe with water-dampened material.


    PCBs have been extracted from water by passing a sample through a
    filter of undecane and Carbowax 400 monostearate supported on
    Chromosorb W (Ahling & Jensen, 1970) or a porous plug of polyurethane
    coated with a suitable gas-liquid chromatographic stationary phase, or
    Amberlite XAD-2 resin (Harvey et al., 1973) followed by elution of the
    PCBs with a solvent. Ahnoff & Josefsson (1974, 1975) have described
    liquid-liquid extraction into cyclohexane.

    Soil and sediment

    In a study by Huckins et al. (1988), sediment samples were thawed at
    room temperature and placed in a hexane-rinsed foil pan and air dried
    for 5 days. The sediment was broken up, homogenized, and mixed with
    anhydrous disodium sulfate until dry, for column extraction. The
    samples were extracted with methylene chloride. PCB residues were
    enriched by adsorption column chromatography on silica gel and
    sulfuric acid silica gel. Prior to GC analysis, nitric acid-rinsed
    copper wool was added to the sediment extract to remove elemental
    sulfur. An aliquot of the PCB residues was diluted in a mixture
    methylene chloride: cyclohexane (1:1) and the bulk of the  o,o-Cl
    substituted PCB components eliminated by eluting the column with
    different solvents. The different PCB congeners were determined by

    The feasibility of cleaning PCB-contaminated soils using a solvent
    extraction method was studied by Reilly et al. (1986). Compared with
    direct incineration of the sludge, the solvent extraction route has a
    number of shortcomings; the detailed design of the extraction plant as
    well as its operation will be quite challenging as an extremely
    leak-tight operation is essential, considering the nature of the
    material handled. Direct incineration will clean the solids much more
    thoroughly than is feasible by solvent extraction under ambient
    conditions. Furthermore, it is inevitable that some residual solvent
    will remain in the solids after processing. The solvent extraction
    process costs essentially the same as direct incineration.

    Biological samples

    Most analysts have used standard methods, developed for organochlorine
    pesticides, in which the PCBs are extracted together with the fat; the
    sample is ground with anhydrous sodium sulfate and extracted with
    petroleum ether or hexane. Porter et al. (1970) studied the optimal
    conditions for this procedure. A dehydrating solvent may be included
    to facilitate the breakdown of cell structures; ethanol (Norn &
    West, 1968) and acetone (Jensen et al., 1973) have been used.

    Reznicek (1987) described a method to extract and determine PCBs in
    blood. The sensitivity of the method was 10 g/litre.  Sample clean-up

    Diverse extraction and clean-up procedures have been devised to
    preferentially remove co-extractives that are present in different
    matrices and interfere with routine quantitative gas chromatographic
    and gas chromatographic-mass spectrometric analysis.

    The analysis of lipid-containing matrices for residues of
    organochlorine pesticides and PCBs is a common procedure. All the
    methods require the separation of the residues from the lipids prior
    to the determination of the PCBs by gas chromatography. The removal of
    the lipids is usually carried out by low-resolution column
    chromatography using an adsorbent, such as silica, alumina, or
    Florisil as the stationary phase. Low-resolution gel permeation
    chromatography has also been used. An electron-capture detector is the
    most commonly used detector, but clean-up procedures may still leave
    electron-capturing species in the extract, so the identities of the
    eluting peaks must be confirmed. In order to overcome some of these
    problems, perchlorination of the PCBs has been used, giving rise to
    one GC peak (decachlorobiphenyl), which is well removed from most
    interfering peaks, but this technique has been found to be
    qualitatively and quantitatively unreliable and unsatisfactory.
    Seymour et al. (1986b,c) attempted to simplify clean-up procedures by
    using high performance liquid chromatography (HPLC) coupled with gas
    chromatography-mass spectroscopy. This latter technique is less
    expensive than it used to be and is the only technique that can
    possibly identify each peak as a PCB before quantification is carried
    out, thereby improving the quality of the result. It is also capable,
    when used in the selective ion monitoring mode (SIM), of detecting
    only PCBs, even in the presence of pesticides, so that sample clean-up
    is further simplified.

    Seymour et al. (1986a) described a clean-up procedure, with a
    preparative, high-performance liquid chromatographic (HPLC) separation
    method for selected pairs of chlorobiphenyl isomers, produced by
    Cadogen coupling in the preparation of individual congeners, to be
    used as standards in congener-specific determination using capillary
    GC methods.

    A routine method for the determination of PCBs in breast milk,
    described by Seymour et al. (1987), is less labour-intensive and more
    cost effective than the traditional methods. These advantages were
    achieved by adsorption of the milk on a polar substrate prior to
    Soxhlet extraction, using a polymeric HPLC column for the clean-up of
    the extract, followed by highly selective capillary GC-MS analysis.

    Methods for the removal of fat from the extract include solvent
    partitioning between hexane and acetonitrile or dimethylformamide, or
    treatment with strong sulfuric acid or ethanolic potassium hydroxide.
    Gel permeation has also been used (Stalling et al., 1972), and Holden
    & Marsden (1969) removed fat on dry, partially deactivated, alumina
    columns. Certain pesticides, such as dieldrin, are destroyed by the
    sulfuric acid treatment, so this method cannot be used if such
    pesticides are to be determined together with PCBs (Jensen et al.,

    Huckins et al. (1988) described the clean-up of fish samples. Tissue
    samples were thawed, mixed, dried with sodium sulfate, and extracted
    in glass columns with methylene chloride. The extract was evaporated
    and the lipid content was determined gravimetrically. Gel permeation
    chromatography was used for removal of lipid from fish sample
    extracts. PCB residues were enriched by adsorption column
    chromatography on silica gel and sulfuric acid silica gel, eluted with
    a mixture of methylene chloride and cyclohexane, and determined by

    PCBs can be separated from organochlorine pesticides by column
    chromatography on Florisil (Mulhern et al., 1971), silica gel (Holden
    & Marsden, 1969; Armour & Burke, 1970; Collins et al., 1972) or on
    charcoal (Berg et al., 1972; Jensen & Sundstrm, 1974a). Several
    laboratories have reported difficulties in repeating results obtained
    by other investigators; the ease of separation appears to depend on
    the characteristics of the absorbent, of the eluting solvent, and of
    the sample extract, though there does not appear to be any difficulty
    in separating all interfering substances, except DDE, a metabolite of
    DDT. Thin-layer chromatography has been used for separation by Norn &
    West (1968), Bagley et al. (1970), and Reinke et al. (1973).

    In many environmental samples, DDE is present in larger amounts than
    the PCBs, and must be removed before their quantitative determination.
    Oxidation procedures have been used to convert DDE to dichlorobenzo-
    phenone; recommended oxidants are potassium dichromate and sulfuric
    acid (West & Norn, 1970b) and chromium (II)oxide and acetic acid
    (Mulhern et al., 1971). Jensen & Sundstrm (1974a), who were
    interested in determining DDT/PCB ratios in environmental samples,
    preferred sodium dichromate in acetic acid with a trace of sulfuric
    acid. They claimed that this does not destroy DDT and its metabolite
    DDD, which may be present in extracts after clean-up with strong
    sulfuric acid, and that using this mixture makes possible the
    quantitative determination of the dichlorobenzophenone from the
    oxidation of DDE.

    Conversion of DDT to DDE can be achieved by treatment with ethanolic
    potassium hydroxide, which also removes interference from elemental
    sulfur (Ahling & Jensen, 1970). Sulfur may also be removed by
    activated Raney nickel (Ahnoff & Josefsson, 1975) or by metallic

    Beck & Mathar (1985) used gel permeation chromatography to clean
    extracts of food of animal origin.

    2.3.2  Separation and identification  Chromatographic separation

    Numerous gas chromatographic studies using packed or capillary columns
    have confirmed the complexity of all commercial PCB formulations. The
    accuracy in determining PCB levels is highly variable and matrix
    dependent. Many factors including: the water solubility, volatility,
    and biodegradability of individual PCBs, will alter the composition of
    a commercial PCB preparation introduced as a pollutant into the
    environment. Thus, the composition of PCB extracts from environmental
    matrices will vary widely and often do not resemble any commercial
    mixture. Quantitative analyses on these mixtures is usually determined
    by pattern- or peak-matching methods, using artificially reconstituted
    mixtures of different commercial formulations. High-resolution, glass
    capillary gas chromatographic analysis can provide a solution.
    Capillary gas chromatography columns, currently in use, are made of
    fused silica, chemically bonded with various stationary phases, to
    achieve a range of different selectivities towards complex samples. In
    general, packed columns have been replaced by capillary columns,
    because of their far superior efficiency. The identities of the
    individual peaks must then be determined by using synthetic standards
    and by retention index addition methods. This latter technique
    predicts the relative retention times (RRT) of specific PCBs and has
    been used to assign the structures of individual PCB congeners. The
    method relies on the RRT values that have been determined for
    synthetic PCB standards. On this basis, Safe et al. (1985a) reported
    the first congener-specific analysis of a PCB preparation and PCBs in
    human milk.

    Some workers use GC with mass selective detection (MSD), which
    quantifies the level of chlorination in a sample extract
    (Alford-Stevens, 1986). Tanabe et al. (1987) and Kannan et al. (1987)
    described a method to determine the 3 toxic, non- ortho-chlorine-
    substituted, coplanar PCBs, 3,4,3',4'-tetra, 3,4,5,3',4'-penta-,
    and 3,4,5,3',4',5'-hexachlorobiphenyl, which are biologically active
    congeners. The method comprised alkali digestion, carbon
    chromatography, and high-resolution gas-chromatography. Using
    this method, it is possible to determine ppt levels of these toxic
    residues in biological samples. Duinker et al. (1988) used
    multidimensional gas chromatography with ECD to determine levels of
    all congeners in some Clophen and Aroclor mixtures and found
    considerable differences between their composition of congeners and
    those in an extract of a seal blubber sample. Using this technique,
    congeners were identified that had, hitherto, been undetected, using
    other analytical techniques. It was possible to identify the toxic
    congeners in the samples studied, even when the relative contribution
    of each congener to the cluster was as low as 0.01%.  Gas-liquid chromatography

    Most analysts use gas-liquid chromatography with an electron-capture
    detector for the separation of PCBs from the extract after clean-up.
    Stationary phases commonly used are silicones or their derivatives,
    for example, DC 200, SF 96, OV 1, and QF 1, or Apiezon L. Jensen &
    Sundstrm (1974a) stated that, with a mixture of SF 96 and QF 1, 14
    peaks could be obtained from Clophen A50, but that Apiezon L gave much
    better resolution. They obtained better peak separation by prior
    fractionation on a charcoal column, which separated the PCBs according
    to the number of  o-chlorine substituents; they regarded such
    refinements as unnecessary in PCB residue analysis, but they may be of
    value in the study of the selective, environmental degradation of
    PCBs. Column temperatures used ranged between 170C and 230C. Glass
    capillary columns are superior to packed columns giving better
    separation of closely-related congeners; they also give good
    separation of PCBs from DDT and its metabolites (Zell et al., 1977;
    Dunn et al., 1984; Beck & Mathar, 1985; Alford-Stevens, 1986; Tanabe
    et al., 1987; Duinker et al., 1988).

    A gas chromatography/electron impact mass spectrometry (GC/EIMS)
    method was used by Erickson et al. (1988) for the determination of
    by-product (non-Aroclor) PCBs. In this method, the recovery of 4
    13C-labelled PCBs was measured to assure adequate recovery of the
    native PCBs from diverse matrices. The complexity of the matrices and
    the high probability of chlorinated organic interferents precluded the
    use of GC/ECD. The best available technique for universal application
    to commercial products, and associated waste, is GC/EIMS. During the
    validation work, the anticipated difficulty of qualitative and
    quantitative data interpretation was confirmed. In addition to the
    inherent problems resulting from extrapolation from 11 standards to
    209 analytes, interpretation of the complex peak clusters is tedious.

    2.3.3  Quantification

    An electron-capture detector (ECD) is the most commonly used
    instrument for the quantification of PCBs. However, the response of
    this detector varies according to the number and location of the
    chlorine atoms in the PCB molecule, resulting in difficulties when the
    sample under investigation contains PCBs that have degraded (Zitko et
    al., 1971).

    Various principles have been used to quantify PCB residues:

    *   comparison of a single peak in the residue with the corresponding
        peak in a commercial reference PCB (Aroclor, Clophen);

    *   comparison of the total response for several peaks in the residue
        with the total response of the corresponding peaks in a reference

    *   comparison of the response of all peaks in the sample with those
        in the reference standard;

    *   perchlorination of PCBs to decachlorobiphenyl followed by
        quantification of this single compound.

    The results obtained using these various methods differ; consequently,
    the precision in these analyses is not very good. Recently, Dunn et
    al. (1984) described a method for the quantification of PCBs using gas
    chromatography data, based on a pattern recognition technique and
    partial least squares in latent variables. The data to which it was
    applied were gas chromatograms of Aroclor 1242, 1248, 1252, and 1260.
    This technique also allows the classification of unknown samples
    (WHO/EURO, 1987).

    Fait et al. (1989) investigated whether the results obtained for total
    PCBs using FSCGC/ECD (see section 2.3), differed significantly from
    those determined using packed column gas chromatography electron
    capture (PCGC/ECD) techniques, within 3 exposure groups. The
    concentrations of individual PCBs were determined in both the serum
    and adipose tissue from 35 transformer repair workers and 17 previous
    repair workers, exposed mainly to Aroclor 1260, in comparison with 56
    non-exposed workers. Eighty-nine PCB peaks were identified. The total
    serum PCBs determined by FSCGC/ECD greatly exceeded that from standard
    PCGC/ECD. The median concentrations in serum were: 43.7, 30.0, and
    16.1 g/litre, and the median concentrations in adipose tissue were:
    3180, 888, and 821 g/kg, respectively. In all workers,
    hexachlorinated and heptachlorinated congeners predominated followed
    by octachlorinated and pentachlorinated species. The 7 major peaks in
    serum and adipose tissue were 2,3,5,6,3',4',5'/ 2,3,4,5,2',4',5'/
    2,3,4,5,2',3',4'-heptachloro-; 2,3,4,2',3',5'-hexachloro-;
    2,4,6,3',4',5'/ 2,4,5,2',4',5'-hexachloro-; 2,3,4,5,2',3',5',6'/
    2,3,4,5,6,2',3',5'-octachloro-; 2,4,5,3',4'/ 3,4,5,2',3'-pentachloro-
    and 2,3,4,2',3',4'/ 2,3,5,6,2',4',5'/ 2,3,4,5,2',4',6'

    The response of the electron capture detector is not equal for all PCB
    components, being much affected by the degree of chlorination, as
    already mentioned (Zitko et al., 1971). This does not lead to
    difficulties when the sample under investigation has been directly
    contaminated by a commercial PCB mixture, as this mixture can be used
    as a standard. Difficulties are encountered when the PCBs in the
    sample have undergone selective environmental degradation. Several
    investigators have noted that the pattern of peaks from such samples
    resembles fairly closely that of one or other of the higher
    chlorinated PCB mixtures, such as Aroclor 1254, and they have compared
    the total area of the peaks with that of the nearest commercial

    product, in order to determine the amount of PCBs in the sample
    (Armour & Burke, 1970; Tuinstra, 1983). Collins et al. (1972) observed
    that, under their conditions, the area of peaks usually encountered in
    extracts of tissue samples was very similar to that of an equivalent
    amount of DDE, thus, DDE could be used for calibration. In order to
    overcome the uncertainties of these procedures, Rote & Murphy (1971)
    divided the peaks into groups according to the number of chlorine
    atoms in the molecule, as determined from mass spectrographic data,
    and calculated the PCB content of each group from the theoretical
    response of the detector to chlorine content. Jensen et al. (1973)
    selected a commercial PCB that included all the peaks from the
    extract; they determined the PCB content of each peak by combined mass
    spectrometry and coulometry, and determined the total PCBs in the
    sample by comparing the height of each peak obtained with the extract
    with those obtained with the reference sample. Simpler methods have
    been used including that of Koeman et al. (1969), who compared the
    height of a single peak, obtained with the extract, with that of a
    peak with the same retention time obtained with a commercial PCB
    mixture, and those of others who averaged out more than one peak for
    this calculation (Reynolds, 1971; Reinke et al., 1973). Rote & Murphy
    (1971) calculated that such procedures may give more than double the
    values obtained by a more accurate method.

    In the characterization of PCB components in PCB mixtures, the
    retention properties of the components of the mixtures, as well as a
    great number of synthesized components, were used to predict a
    complete analysis of mixtures as Aroclors 1242, 1254, and 1260. Jensen
    & Sundstrm (1974a) synthesized a large number of reference substances
    and were able to identify almost 60 components in Clophen A50 and A60.

    Attempting to account for unidentified peaks, authors have used the
    chromatographic retention indices of available components to calculate
    such data for missing ones. The identity of many peaks could not,
    however, be determined unambiguously. Some of these uncertainties have
    been resolved by the application of techniques other than the
    comparison of retention times e.g., MS, NMR, and IR. The efficiency of
    packed columns in GLC is not sufficient to allow their use for the
    accurate analysis of complex mixtures, in most cases. Another approach
    to the use of packed columns involves the use of columns with various
    selectivities. In this way, complete analysis of all components in
    Aroclors has been claimed with the use of up to 12 columns. The
    strongly increased GLC separation offered by capillary columns has
    been used to advantage in the analysis of technical formulations, in
    some cases the eluate was analysed by MS. To identify individual
    congeners, gas-liquid (using glass capillaries with different
    coatings) chromatography (GLC) was used by Albro & Parker (1979) and
    Albro et al. (1981). Hydrogen flame ionization detection (HFID) and
    electron capture detection (ECD) and MS were used by Duinker &
    Hillebrand (1983).

    2.3.4  Accuracy of PCB determinations

    A group of 8 analysts, engaged in an investigation of pollution in the
    North Sea, undertook a collaborative study to determine the PCB
    content of a sample of fish oil, using the methods currently employed
    in their laboratories (International Council for the Exploration of
    the Sea, 1974). The PCB values obtained ranged from 1.0 to 3.9 mg/kg
    with a mean of 1.97 mg/kg and a standard deviation of 0.93 mg/kg.
    Better agreement was obtained with the same fish oil fortified with
    PCBs at a concentration of 10 mg/kg; the mean of the results for the
    fortified sample was 10.0 mg/kg with a standard deviation of
    1.1 mg/kg.

    A probable source of error is incomplete initial extraction of PCBs
    from a sample (Holden & Marsden, 1969). Another source of variation
    between laboratories lies in the method used to quantify gas-liquid
    chromatographic peaks; Van Hove Holdrinet (1975) considered this to be
    the major source of error.

    It is evident that caution should be exercised in accepting the
    analytical results from a laboratory, particularly for samples with a
    low PCB content, until the competence of the laboratory has been
    established by an inter-laboratory collaborative study (Tuinstra,

    Schulte & Malisch (1984) described a method to determine the real PCB
    contents of environmental samples. A technical PCB mixture of known
    composition was used for calibration. The PCB concentrations were
    determined in samples of human milk and butter and the calculated
    contents were 50% and 40% lower, respectively, than the values
    obtained by the usual calculation based on evaluation of some higher
    peaks of technical PCB mixtures.

    2.3.5  Confirmation

    Since Jensen first identified as PCBs hitherto unknown substances that
    interfered in the glass-liquid chromatographic determination of
    organochlorine pesticides using mass spectrographic data, other
    investigators have confirmed the presence of PCBs in environmental
    samples by combining gas-liquid chromatography with mass spectrometry
    (Bagley et al., 1970) and with coulometry, to measure the chlorine
    content. The conversion of PCBs to bicyclohexyl and decachlorobiphenyl
    is further confirmation (Berg et al., 1972). The widespread
    distribution of PCBs is now well established, and, as adequate methods
    are available to remove interference from organochlorine pesticides,

    there is no evidence of the presence of other interfering substances
    in the types of sample that have so far been analysed, down to a limit
    of detection of around 0.01 mg/kg. This does not necessarily apply to
    other types of sample, particularly when very low levels are being
    sought; Ahnoff & Josefsson (1973) reported a number of unknown
    interfering substances, when measuring PCBs in water at levels below
    1 ng/litre. One of these substances was subsequently identified as
    elemental sulfur. They recommend confirmation by mass fragmentography
    for such samples.

    2.3.6  Detection limits

    The limits of determination using low or high resolution mass
    spectrometry are 0.01-1 pg per injection of each congener. The
    detection levels in samples depend on the sample size and matrix.
    Using an air sampling device described by Rappe et al. (1985b), a
    detection level of 0.05 pg/m3 per congener could be determined in
    ambient air (WHO/EURO, 1987).

    In general, other substances are not considered to interfere at levels
    of about 0.01 mg/kg. In river water and air, levels of 1 ng/litre and
    0.3 ng/m3, respectively, are reported to be the detection limits of
    PCBs (WHO/EURO, 1987). Tuinstra (1983) found a limit of detection for
    individual chlorobiphenyls in environmental and biological samples, of
    less than 1 g/kg (see Table 7).

    The results for sewage sludge, eel, grass, cow's milk, and human fat
    are given in Table 7 (Tuinstra, 1983). Individual chlorobiphenyls were
    also estimated in the monitoring programme for environmental and
    biological samples in the Netherlands.

    2.4  Codex questionnaire on analytical methods

    2.4.1  Interpretation and comparability of data

    Monitoring data are available from many sources in many countries.
    They have been obtained using various methodologies, such as different
    sampling techniques and different methods of analysis and
    quantification. Limits of determination reported vary by a factor of
    1000 or more.

    Given this situation, data on levels of PCBs have to be interpreted
    with the greatest care. Comparisons can only be made between data from
    the same laboratory, using the same validated technique over a long
    period. Comparisons between data from different laboratories have to
    be limited to the very few cases, where very strict inter-laboratory
    checks have been made on the basis of the same sampling and analytical
    techniques. Indications about trends can only be obtained when taking
    into account these basic considerations (Beck & Mathar, 1985; Tuinstra
    et al., 1985b,c).

    In June 1985, a questionnaire was distributed to all Codex Contact
    Points with the aim of providing background information on PCBs for
    the ad hoc working group on contaminants to compare such factors as
    methods of analysis, quantification, monitoring, etc. Eighteen out of
    22 countries responded to the questionnaire.

    In some cases, the information given was incomplete, but it is
    apparent that a variety of clean-up methods is employed. Where good
    laboratory practices are followed and tests indicate close to 100%
    recovery of standards from spiked samples, the main effect of
    different clean-up procedures will be on the limit of detection.

    For gas chromatography, 6 countries reported that they used capillary
    columns as alternative or confirmatory systems. Among the respondents,
    the Netherlands and the Federal Republic of Germany routinely used
    capillary columns and specific PCB isomers as regulatory standards.
    The types of packed column materials used varied considerably. With
    respect to quantification, pattern comparison with standards of
    various PCB formulations was the method most favoured, though some
    countries specified the use of certain combinations of peaks. In
    several cases, the methods being used were stated to have been
    collaboratively tested, or checked by inter-laboratory ring tests.

    During the sixties, packed column chromatography was the most widely
    used method in the determination of PCBs. Results obtained with this
    technique varied widely between laboratories, and were much influenced
    by the method of quantification chosen and by the PCB mixture used as
    a standard. Chemical conversion methods, especially perchlorination,
    have also been used. These methods are quite sensitive, but do not
    allow for peak pattern identification. Another drawback of
    perchlorination is that conversion of less chlorinated biphenyls is
    not quantitative.

    Sensitivity is sufficient, if adequate clean-up methods are used.
    Combined gas chromatography/mass spectrometry has a somewhat lower
    sensitivity, needs more expensive equipment, and is not considered
    suitable for routine work. The results obtained using these techniques
    may vary widely and most of them can only be used as rough estimates.

    When capillary columns are used with temperature programming, almost
    all PCB isomers and congeners normally present in samples can be
    identified. This method is now considered to be the best available
    technique. However, it is important to decide which isomers should be
    used as guiding substances.

        Table 7.  Typical values of individual chlorobiphenyls in Dutch environmental and
              biological samples. Peak numbering according to IUPAC rulesa

    PCB      Structure           Sewage     Eel         Grass      Cow's        Human
             compound            sludge                            milk         fat
                                 g/kg      g/kg       g/kg      g/kg        g/kg
                                 (dm)b      product     (dm)b,c    fatc         fat

     28d     2,4,4'                60         35        -c          -c           45
     52d     2,5,2'5'              22        110        0.4          2.1         10
     44      2,3,2'5'              20         34        0.2          0.9         10
     95      2,3,6,2'5'            58        130        0.7          1.6         30
    101d     2,4,5,2'5'            30         85        0.6          3.1         15
    151      2,3,5,6,2'5'           9         24        0.2          0.6         10
    149      2,3,6,2'4'5'          42         90        0.6          2.5         15
    118      2,4,5,3'4'            20        110        0.3         -c           80
    153d     2,4,5,2'4'5'          54        180        0.7         13          295
    141      2,3,4,5,2'5'          10         40        0.2          0.6         <5
    138d     2,3,4,2'4'5'          45        200        0.7         11          235
    128      2,3,4,2'3'4'           7         20       <0.1          1.2         15
    180d     2,3,4,5,2'4'5'        33         80        0.5          6.4        205
    170      2,3,4,5,2'3'4'        10         30        0.2          1.8         90
    201      2,3,4,5,2'3'5'6'      <5         10       <0.1         <0.5         20

    a  From: Tuinstra (1983).
    b  dm = dry matter.
    c  nd = not determined.
    d  Monitoring compounds.

    2.5  Activities of the WHO Regional Office for Europe

    The WHO Regional Office for Europe (WHO/EURO) has an ongoing programme
    related to PCBs, as well as to other chlorinated hydrocarbons,
    including polychlorinated- para-dibenzodioxins (PCDDs) and
    polychlorinated dibenzofurans (PCDFs). Within this programme,
    practical guidelines to prevent and control accidental and
    environmental exposures to these chemicals have been published in the
    Environmental Health Series of WHO/EURO (1987). The other important
    project within this programme dealt with the assessment of the health
    risks to infants associated with contamination of mother's milk. This
    assessment was completed by a WHO/EURO Expert Consultation held in
    Abano Terme, Italy, in 1987, and the output of this consultation has

    been published in the Environmental Health Series of WHO/EURO (1988).
    In order to produce more data on exposure levels through human milk,
    WHO/EURO has been coordinating analytical field studies in which
    several countries have participated. The results of these studies have
    been published in the Environmental Health Series of WHO/EURO (1989).
    This document also includes the results of interlaboratory quality
    control studies on levels of PCBs, PCDFs, and PCDDs in human milk. In
    the first series of studies, 12 laboratories were involved. The second
    round of the quality control studies has been completed, with the
    participation of additional laboratories, and the results will be
    published. Furthermore, the repetition of the analytical field studies
    on the levels of PCBs, PCDFs, and PCDDs in human milk will be
    implemented in 1991 and coordinated at WHO/EURO.

    2.6  Appraisal

    Since the congener composition and relative concentrations of the
    individual components in PCB extracts from environmental and
    biological samples differ markedly from those in commercial PCB
    mixtures, the quantitative determination of the PCB contents of such
    samples presents a special problem. Various approaches to the
    quantitative determination of PCBs have been reported including:
    attempts to determine the total PCB concentration through
    perchlorination of the mixture; identification of selected
    chromatographic peaks through gas chromatographic techniques with
    packed columns using certain commercial products as standards; as well
    as attempts to carry out congener-specific analysis, based on high
    resolution chromatographic separation followed by identification and
    quantification by mass spectrometry using synthetic standards. This
    last method is considered the best at present, though it is not
    feasible for all laboratories. Although the concentration values
    obtained from the various methods might be similar, such comparison
    will be limited and is of questionable value for most purposes. The
    occurrence of specific PCB congeners in various samples and a
    consideration of the relative toxicity and persistence of the
    congeners have been suggested as a basis for a congener-specific
    analytical approach. While this approach can be useful, particularly
    in risk/hazard assessment exercises, it must be realized that it is
    based on the present knowledge about the occurrence, persistence, and
    toxicity of specific congeners. It does not take into consideration
    potentially unrecognized toxicities associated with the same or
    different congeners, which may be present in a sample, also it is not
    feasible in some countries. Therefore, further research in this area
    should continue to improve the basis for monitoring programmes and for
    a congener-specific approach.

    In the selection of areas with high levels of contamination, in order
    to establish priorities for action, it is considered that analytical
    competence and the use of adequate controls and standards is more
    important than highly accurate laboratory analysis. Also, the quality
    and usefulness of analytical data depend critically on the validity of
    the samples and the adequacy of the sampling programme. A quality
    assurance programme and collaborative studies should be part of any
    long-term study on PCBs, since there are several possible sources of
    error. In this situation, data on levels of PCBs have to be
    interpreted with the greatest care and, in general, definitive
    comparison can only be made between data from laboratories using the
    same techniques and interpretation of results.


    3.1  Natural occurrence

    Polychlorinated biphenyls are aromatic chemicals that do not occur
    naturally in the environment.

    3.2  Man-made sources

    3.2.1  Production levels and processes, uses

    The first chlorinated biphenyl was synthesized in 1864, but it was not
    until 1929/1930 that the PCBs were produced commercially for use:

     (a) as dielectrics in transformers and large capacitors;

     (b) in heat transfer and hydraulic systems;

     (c) in the formulation of lubricating and cutting oils and wax

     (d) as plasticizers in paints, and as ink solvent/carriers in
    carbonless copy paper, adhesives, sealants, flame retardants, and
    plastics (Hutzinger et al., 1974; Pomerantz et al., 1978).

    An extensive review of the uses of PCBs is given in DFG (1988).  World production figures

    Over one million tonnes of PCBs have been produced commercially under
    a number of trade names, such as Aroclor, Fenchlor, Clophen, and

    Details of the production and uses of PCBs in the USA have been
    released, and have been summarized by Nisbet & Sarofim (1972). Annual
    production increased steadily from 1930 and reached a maximum in 1970
    of 33 000 tonnes. Of this, 56% was used as a dielectric (36% in
    capacitors and 20% in transformers). Various plasticizer outlets
    accounted for 30%, hydraulic fluids and lubricants, 12%, and heat
    transfer liquids, 1.5%. During this peak year, 65% of the production
    was of the 42% chlorinated type, 25% was less chlorinated, and the
    remainder more chlorinated. After 1970, production decreased sharply
    owing to the voluntary limitation of sales by the Monsanto Company,
    the major manufacturer in the USA.

    Following the restriction of sales for dissipative uses, the
    percentage of PCBs sold as dielectrics rose to 77% in 1971 and the
    proportion of highly chlorinated products was considerably reduced;
    Aroclor 1016 replaced Aroclor 1242. In Japan, 44 800 tonnes of PCBs
    were used from 1962 to 1971; of this, 65.4% was used in the electrical
    industry, 11.3% in heat exchangers, 17.9% in carbonless copying paper,
    and 5.4% for other dissipative uses (Ishi, 1972).

    During the period 1980-84, the production in EEC member states was as
    follows: France, 16 200; Federal Republic of Germany, 24 200; Italy,
    4500; and Spain, 3400 tonnes. After 1984, production was continued
    only in France and Spain (Bletchly, 1985; WHO/EURO, 1987).

    By the end of 1980, the total amount of PCBs produced was 1 054 800
    tonnes (of which approximately half was used in transformers and
    capacitors, see Table 8), divided between the following countries (in
    tonnes): USA, 647 700; Federal Republic of Germany, 130 800; France,
    101 600; United Kingdom, 66 800; Japan, 59 300; Spain, 25 100; and
    Italy, 23 500 (Bletchly, 1983).

    In addition, Czechoslovakia and the USSR have manufactured PCBs for
    their domestic market under the trade names of Delor and Sovol,
    respectively, but the data on production quantities are not available.

    According to an OECD report, transformers and capacitors provided the
    major outlets for PCBs in most OECD countries in 1971. In 1972,
    several countries restricted sales; in Sweden the importation and use
    of PCBs were restricted by law; in the United Kingdom, as in the USA,
    sales were voluntarily restricted to the lower chlorinated PCBs for
    use as dielectrics in enclosed systems, and, in the USA in 1979,
    manufacture, use, handling, storage, and disposal were promulgated. As
    late as 1985, a final rule concerning the restriction and conditions
    on the use of PCB transformers was published (USEPA, 1985). In Japan,
    the production and use of PCBs were banned in 1972.

    The 24 OECD countries adopted a Decision in 1973, limiting the use of
    PCBs to certain specific applications and asking for the control of
    the manufacture, import, and export of bulk PCBs, for adequate waste
    treatment and for a special labelling system for PCBs and
    PCB-containing products. On 13 February 1987, the Council of the
    Organization for Economic Co-operation and Development (OECD) adopted
    a further Decision-Recommendation (C(87)2(final)) on "Further measures
    for the protection of the environment by control of polychlorinated

        Table 8.  Estimated usage of PCBs in transformers and large capacitors in
              a number of OECD countries in 1930-80 (in tonnes)a

    Country            Usage in          Usage in         Total
                       transformers      capacitors

    France               50 700            8 800          59 500
    Federal Republic     44 400           17 700          62 100
    of Germany
    Italy                10 400            1 500          11 900
    Japan                37 200b          37 200
    Spain                20 100            3 400          23 500
    United Kingdom        5 800            8 100          13 900
    United States       125 800          130 400         256 200
    of America

    Total               294 400          169 900         464 300

    a  From: WHO/EURO (1987).
    b  Includes the usage in both transformers and capacitors

    biphenyls". With this Decision-Recommendation, the OECD Member
    countries committed themselves to ban virtually all new uses of PCBs,
    accelerate the phasing out of PCBs from existing uses, control PCBs in
    contaminated products, articles, or equipment, and ensure appropriate
    disposal methods for PCB-containing waste. The uses of PCBs have been
    virtually restricted to those in "closed systems". In 1976, an EEG
    Directive made the limitations of the use compulsory for the EEG
    Member States. Other Directives, such as those on waste treatment and
    disposal, followed (van der Kolk, 1984a, Personal communication).  Manufacturing processes

    Industrial manufacturing of PCBs is based on the chlorination of
    biphenyl by anhydrous chlorine, under heated reaction conditions and
    in the presence of suitable catalysts (e.g., iron-chloride). Depending
    on the reaction conditions, a degree of chlorination varying between
    21% and 68% (weight percentage, w/w) can be achieved.

    The yield is always a mixture of different compounds and congeners.
    Commercial mixtures generally have been purified by filtration and
    fractional distillation, but, in spite of this, they have been found
    to contain many impurities (WHO/EURO, 1987). In general, commercial
    PCB products contain impurities, mainly polychlorinated dibenzofurans

    Rappe et al. (1985d) cf. WHO/EURO (1987) analysed a series of
    commercial PCBs, using a new clean-up technique based on reverse-phase
    chromatography on a carbon column followed by a fluorosil column. In
    all PCB products, PCDFs were found at levels varying from a few mg/kg
    up to 40 mg/kg. The chlorination pattern of the PCDFs was found to
    vary with the chlorination level of the PCBs. In most products,
    2,3,7,8-substituted tetra-, penta-, and hexa-CDFs were the major

    3.2.2  Uses

    PCBs have been widely used in electrical equipment, such as capacitors
    and transformers. These have often been considered to be closed
    systems, though small amounts of PCBs can frequently be found on the
    outer metal surface of such equipment.

    Smaller volumes of PCBs have often been used as fire-resistant liquid
    in nominally closed systems, such as hydraulic and heat exchange
    systems (WHO/EURO, 1988).

    Broadhurst (1972) reviewed the many technical applications of PCBs
    that appear in the literature and in patent specifications, and
    indicate the possibility of a widespread, non-occupational, low-level
    exposure to PCBs, other than that derived from the diet. PCBs are used
    in the home in ballast capacitors for fluorescent lighting, and
    exposure from pressure-sensitive copying paper has not been limited to
    office workers. The valuable properties of PCBs as plasticizers has
    led to their use in furnishings, interior decoration, and building
    construction; examples are surface treatment for textiles, adhesive
    for waterproof wall coatings, paints, and sealant putties. PCBs have
    been used as plasticizers for plastic materials and in the formulation
    of printing inks.

    The value of PCBs for industrial applications depends on their
    chemical inertness, resistance to heat, non-flammability, low vapour
    pressure (particularly with the higher chlorinated compounds), and
    high dielectric constant.

    Data on the usage of technical PCB mixtures in Europe are scarce. In
    the 1960s and early 1970s, PCBs were used in (WHO/EURO, 1987):

     (a) completely closed systems;
     (b) nominally closed systems;
     (c) open-ended applications.  Completely closed systems

    PCBs have been widely used in electrical equipment, such as capacitors
    and transformers, which are considered to be completely closed
    systems. Historically, capacitors are the single largest PCB-use
    category. The PCB mixtures used for this purpose are, for example,
    Pyralene 3010, Aroclor 1016, 1221, and, earlier, also Aroclor 1242 and
    1254. The amounts used in a number of OECD countries are presented in
    Table 8 (OECD, 1982; Bletchly, 1983; Callahan et al., 1983).

    Since the late 1970s and the beginning of the 1980s, PCB-filled
    capacitors have largely been superseded by capacitors with a non-PCB
    dielectric fluid. The tendency for this substitution varies from
    country to country, for example, it started in Sweden and Finland in
    1982, and in Norway in 1985.

    The technical PCB mixtures used in transformers are mostly highly
    chlorinated like Aroclor 1254 and 1260. In general, the PCBs are used
    in combination with tri- and tetrachlorobenzenes as mixtures called

    The amounts of PCBs used in transformers differ in different
    countries. In France, where most transformers are placed indoors, the
    major dielectric fluid is PCBs or  Askarels, which are both flame
    retardants, while in Scandinavia, where most capacitors are placed
    outdoors, mineral oils (with a lower melting point) are frequently

    During the 1980s, there has been a marked interest in replacing the
    PCBs, mainly in indoor transformers, as a result of serious accidents,
    for example, in Binghamton, San Francisco, Miami in the USA, and Reims
    in France. Various products are used for this exchange, such as
    mineral oils, silicone oils, perchloroethylene, and other chlorinated
    products (WHO/EURO, 1987).  Nominally closed systems

    Smaller volumes of PCBs have frequently been used as fire-resistant
    liquid in nominally closed systems, such as hydraulic and heat
    transfer exchange systems (for example, trade names Pydraul and
    Therminol FR, containing Aroclor 1242, 1248, 1254, and 1260). PCBs are
    used as a working fluid in vacuum pumps (Aroclor 1248, 1254), which
    can also be considered as nominally closed systems (WHO/EURO, 1987).  Open-ended applications

    With open-ended applications of PCB, both the emissions into the
    environment and the levels of occupational exposure are more
    pronounced. The major open-ended applications include use as a
    plasticizer (in PVC, neoprene, and other artificial chlorinated
    rubbers). Other open-ended uses, such as surface coatings, paints,
    inks, adhesives, pesticide extenders, microencapsulation of dyes, and
    carbonless copy paper contribute smaller volumes into the environment.
    PCBs have also been used in immersion oils for microscopes, as
    catalysts in the chemical industry, in casting waxes in the iron/steel
    industry (decachlorobiphenyl), and in cutting and lubricating oils
    (WHO/EURO, 1987).  Contamination of other compounds

    In addition to the above uses of PCBs, numerous halogenated compounds
    may contain PCBs in small amounts as a contaminant (US EPA, 1983).

    3.2.3  Loss into the environment

    PCBs are dispersed into the environment through atmospheric transport
    and, on a more regional scale, following release into water. PCBs are
    also mobilized in the soil or landfills, but the rates of dispersion
    and subsequent transfer to biota and humans are difficult to estimate.

    More highly chlorinated forms become most prevalent in compartments
    further along the pathway chains. The analytical methods used to
    quantify PCBs in the environment and biota vary greatly within, and
    between, countries. Thus, comparisons can only be made in a very broad
    sense and could, to some extent, be erroneous (WHO/EURO, 1988).

    An overview of prevention and control measures of accidental and
    environmental exposures is given in WHO/EURO (1987).  Routes of environmental pollution

    Surveys of the sources of environmental pollution with PCBs were made
    before production and use became limited, and the information
    available may not now apply in North America and elsewhere. Only 20%
    of the annual production in the USA can be regarded as a net increase
    in current usage, and the remainder is balanced by a loss to the
    environment. More than one-half of this entered dumps and landfills
    and it has been calculated that 0.3 million tonnes of PCBs have
    accumulated in such locations in North America, since 1930 (Nisbet &
    Sarofim, 1972). Much of this was originally enclosed in containers,
    such as capacitors, or was in plasticized resins and will not be
    released until the containing medium decays. The diffusion of PCBs
    from landfills is likely to be slow, on account of their low
    volatility and low water solubility. Carnes et al. (1973) found little
    leaching from the one site that they tested.

    The concentration of PCBs in emissions from several municipal sanitary
    landfills and refuse and sewage sludge incinerators were determined in
    the Midwest of the USA. Sanitary landfills continuously emit the
    gaseous products of anaerobic fermentation together with other
    volatile materials into the atmosphere. A projection, based on the
    amount of methane generated annually from landfills and a PCB to
    methane ratio of 0.3 g PCBs/m3 of methane found from the landfills
    sampled, indicates that the annual PCB emissions from sanitary
    landfills in the USA are of the order of 10-100 kg/year. The
    concentrations of PCBs from the incinerator stacks ranged from
    0.3-3 g/m3 and the annual emissions per stack were 0.25 kg/year.
    These estimates are very small in comparison with the 900 000 kg
    PCBs/year estimated to cycle through the atmosphere over the USA,
    annually (Murphy et al., 1985).

    Scrap transformer fluid containing PCBs has been used in the USA in
    amounts of about 10 tonnes/year in pesticide formulations (Panel on
    Hazardous Trace Substances, 1972, cf. WHO/EURO, 1988), and this
    unauthorized use has led to the local contamination of milk supplies.

    Pressure sensitive duplicating paper (carbonless copying paper)
    containing PCBs has found its way into waste paper supplies and has
    been recycled into paper and board used as food packaging materials,
    but not since 1970; paints for coating the bottom of ships contained
    3-5% of PCBs, about 3% of the annual quantity imported into Sweden has
    been used for this purpose, and this has been a source of plankton
    contamination (Jensen et al., 1972a).

    Schecter (1987) described the contamination of drinking-water by the
    use of submersible water pumps which, in certain instances, contained
    PCBs in the oil. When the pumps leak, PCBs may be released into the

    In addition, the US EPA, in 1980, estimated that over 1 000 000 wells
    in the USA may have PCB capacitors in the well motors. Levels recorded
    in drinking-water range from 0.26 to 57 g/litre compared with
    1 g/litre considered safe in the guidelines for New York State. The
    oil from these pumps contained 630 000-24 000 000 g/kg of PCBs.

    Stehr et al. (1985) studied the possibility of contamination with PCBs
    of oils and oil-filled devices used by amateur radio operators. Two of
    77 oil samples contained more than 50 mg/kg.  Release of PCBs into the atmosphere

    There appears to be little atmospheric contamination during the
    manufacture and processing of PCBs, but this can occur during their
    subsequent use and disposal. Although PCBs have a low volatility,
    there may be an appreciable loss to the atmosphere during the lifetime
    of a PCB-plasticized resin, particularly of the lower chlorinated
    products. Further pollution may occur during the incineration of
    industrial and municipal waste. Most municipal incinerators are not
    very effective in destroying PCBs; efficient incinerators can be
    designed for this purpose (Oehme et al., 1987), though the higher
    chlorinated PCBs are more resistant to pyrolysis. Secondary sources of
    atmospheric pollution are volatilization from soil, and the drying of
    sewage sludge. Furthermore, there is evidence that, even at ambient
    temperatures, PCBs will enter the atmosphere by volatilization from
    soils and water bodies, landfill sites etc. (section 4.1.1).  Leakage and disposal of PCBs in industry

    Eschenroeder et al. (1986) analysed PCB risks using estimates of human
    intake of PCBs originating from accidental spills from electrical
    equipment. Equipment spills without controls resulted in a human
    intake of PCBs of, at the most, 2 ng/day via the water exposure
    pathway. This was negligible in comparison with the intakes calculated
    on the basis of fish consumption. The inhalation exposure of
    approximately 100 persons living in the vicinity of a spill in
    Southern California was determined to equal the PCB intakes of a
    fish-eating population.

    3.2.4  Thermal decomposition of PCBs

    It has been found by Buser et al. (1978a,b) that PCBs can be converted
    to PCDFs under pyrolytic conditions. The pyrolysis of a commercial PCB
    mixture in a sealed quartz ampoule, in the presence of air, yielded a
    mixture including about 30 major and more than 30 minor PCDF

    Buser & Rappe (1979) studied the pyrolysis (at 600C) of 15 individual
    PCB isomers and demonstrated the presence of PCDFs via intramolecular
    cyclizations, where m + n varies from 4 to 8 (Fig. 1). The
    thermochemical generation of PCDFs from PCBs was found to follow 4
    general reaction routes including loss of  ortho-Cl; loss of HCl
    involving a 2,3-chlorine shift at the benzene nucleus; loss of
     ortho-HCl and loss of  ortho-H (Buser, 1985; Hutzinger et al.,

    FIGURE 1

    The maximum yield of PCDFs was about 10%, calculated on the amount of
    PCBs decomposed, and the optimal temperature was between 550 and
    650/700C (Bentley, 1983). Thus, the uncontrolled burning of PCBs can
    be an important occupational and environmental source of toxic and
    hazardous PCDFs and it is recommended that all destruction of
    PCB-contaminated waste should be carefully controlled, especially with
    regard to the burning temperature (above 1000C), residence time, and
    turbulence (Bentley, 1983; WHO/EURO, 1987).

    In the temperature range 300-400C, Morita et al. (1978) reported that
    the yield of conversion seemed to be in the mg/kg range. However,
    Nagayama et al. (1981) reported a dramatic increase in the levels of
    PCDFs at these rather low temperatures, in the presence of stainless
    steel or nickel.

    No, or very low levels of, PCDDs have been reported from the pyrolysis
    of PCBs. However, pyrolysis of a mixture of PCBs and chlorobenzenes
    (product  Askarel) can yield both PCDFs and PCDDs (Buser, 1979).

    Rappe et al. (1985b) found that various types of industrial
    incinerators, such as copper smelters and steel mills generate PCDFs
    and PCDDs. Pyrolysis of chlorinated polymers like polyvinylchloride
    (PVC) and Saran also generate these compounds and exhaust gases of
    motor cars and their motor oil may contain PCDDs and PCDFs (WHO/EURO,

    In a State Office Building in the centre of Binghamton, New York, a
    fire, in conjunction with several explosions, occurred in the basement
    mechanical room, in 1981. Approximately 750 litres of  Askarel, a
    dielectric fluid composed of 65% PCBs (Aroclor 1254) and 35%
    polychlorinated benzenes, leaked from a transformer and caught fire.
    Pyrolysis of the  Askarel led to the formation of a fine oily soot
    that spread throughout the building via 2 ventilation shafts. Samples
    taken several days after the fire showed average concentrations of
    PCBs in the air of the building of 1.5 g/m3. The average result for
    surfaces ranged from 4.6 to 162.2 g/m2. TCDFs and PCDDs were also
    present. The soot samples were analysed for pyrolysis products. They
    contained average levels of 3 mg TCDD/kg and 199 mg 2,3,7,8-TCDF/kg
    (Fitzgerald et al., 1989). Achilles (1983) reported the following
    levels in the deposited smut; 2160 mg PCDFs/kg and 20 mg PCDDs/kg
    (including 0.6 mg 2,3,7,8-TCDD/kg).

    In the soot from the Binghamton, Reims, and Stockholm accidents, high
    levels of polychlorinated biphenylenes (PCBPs) were identified as well
    as the PCDFs (Fig. 2) (Rappe et al., 1982, 1985).

    Between 1981 and 1985, a number of accidents in electrical equipment
    were reported from different countries; 28 accidents were mentioned in
    WHO/EURO (1987) including actual capacitor explosions, capacitor
    fires, and transformer accidents. In all eases, the accident site was
    contaminated by PCDFs, average levels of total PCDFs being in the
    range of 1-5 g/m2.

    FIGURE 2

    Hutzinger et al. (1985) also mentioned the presence of polychlorinated
    pyrenes (PCPYs).

    In the period 1977-85, particulates and flue gas from municipal
    incinerators and hazardous waste incinerators in Canada, Denmark,
    Netherlands, Sweden, and Switzerland were investigated. It was found
    that emissions from incinerators contained many different PCDF and
    PCDD isomers. The total levels ranged from ng/m3 to g/m3. Fly-ash
    contained levels of 0.1-0.6 mg/kg (Buser & Bosshardt, 1978; Rappe et
    al., 1985c; WHO/EURO, 1987).

    Rappe et al. (1985b) studied the emissions of the municipal solid
    waste incinerator in Umea, Sweden. The levels of PCDDs and PCDFs
    varied under different burning conditions. The amount of dioxins
    formed seems to be dependent on the chlorine content in the waste, as
    well as the construction of the incinerator. The critical parameters
    seem to be temperature, residence time, turbulence, and excess air

    The 2,3,7,8-tetra-CDD was always found to be a very minor constituent,
    whereas the 1,2,3,7,8-penta-CDD in all samples gave a medium-sized
    peak. The 2,3,7,8-substituted PCDFs were always middle or major
    components (WHO/EURO, 1987).

    The fact that PCBs may be thermally converted to PCDFs has raised
    concern that similar conversions might occur in electrical equipment,
    such as capacitors and transformers, in which the dielectric fluids
    used are subjected to modest temperature rises accompanied by
    electrical stress. Brown et al. (1988) investigated the presence of
    PCDFs in both used and unused capacitors and transformers and did not
    find any evidence of an increase in PCDFs levels in the heavily used
    capacitor or the transformer PCBs compared with levels in unused

    For a number of years, concern has been expressed regarding the
    release of PCBs and other dangerous compounds when fluorescent light
    ballasts "burn out". The breakdown products may contain vapours and
    condensed particles of PCBs and asphalt. In response to concern at a
    school, the US EPA met with officials of Blaine Elementary School,
    because of material leaking from some fluorescent light fixtures. It
    was determined that the leaking material ("oil") contained PCBs
    (Aroclor 1242 or 1260). Air samples collected following the burn out
    of such lights, at different distances from the light fixture, gave
    concentrations of 0.166 and 0.012 mg/m3, respectively, 1 and 6 m from
    the light. Three days later, levels of 0.004-0.001 mg/m3 were still
    found. In a second series of tests, both burn-out and non-burn-out
    ballasts were heated to 150C, 300C, and 400C, in a chamber. No PCBs
    were detected at 150C. At 300C, concentrations ranged from 0.55 to
    1.70 mg/m3 and, at 400C, 2.54 to 28.2 mg/m3. Wipe samples were
    taken in schoolrooms after burn-outs; average concentrations of
    Aroclor of 0.34 and 1.22 g/cm2 were found. It is obvious that PCBs
    and asphalt contamination, both surface and atmospheric, can occur
    when fluorescent lamp ballasts burn out.

    The most serious potential contamination results when thermal runaway
    takes place. Thermal runaway volatilizes the asphalt potting compound
    and may rupture the capacitor. When the potting compound and the PCBs
    are exposed to high temperatures, some of both materials vapourizes.
    As the vapours pass through the atmosphere they condense into freely
    divided aerosols, less than 1 m in diameter. Much of the visible
    fumes results from volatilization of the asphalt (Anon., 1987).


    4.1  Transport and distribution between media

    A more detailed review of transport mechanisms can be found in Jury et
    al. (1987).

    4.1.1  Transport in air

    The virtually universal distribution of PCBs throughout the world,
    including the arctic and other remote areas, suggests that PCBs are
    transported in air (Risebrough & de Lappe, 1972). The ability of PCBs
    to co-distill, volatilize from landfills into the atmosphere
    (adsorption to aerosols with particle size of less than 0.05-20 m),
    and resist degradation at low incinerating temperatures, makes
    atmospheric transport the primary mode of global distribution within
    the troposphere and stratosphere (Nisbet & Sarofim, 1972; Eisenreich
    et al., 1981). PCBs have been measured in air samples at Eniwetok
    Atoll in the North Pacific Ocean (Atlas & Giam, 1981), over the North
    Atlantic (Giam et al., 1978), and in the Gulf of Mexico (Giam et al.,
    1978, 1980). Murphy et al. (1985) estimated that approximately
    18 000 kg of PCBs are present in the atmosphere over the USA, at any
    given time. The authors also estimated that, if these PCBs had an
    atmospheric residence time of one week, then about 900 000 kg/year of
    PCBs cycle through the atmosphere of the USA.

    Nisbet & Sarofim (1972) suggested that most of the airborne PCBs will
    be adsorbed on any particles present. The half-life of particles in
    the air will depend greatly on the size of the particles and the
    extent of atmospheric precipitation. Most will be deposited within 2-3
    days in their areas of origin (usually urban), the small amount
    attached to fine particles will last in the atmosphere for longer
    periods and can be transported to more remote regions.

    Sdergren (1972) collected airborne fallout in southern Sweden and
    found regional differences in PCB levels, with mean monthly levels
    ranging from 620 ng/m2 per month to 10 510 ng/m2 per month. The
    lowest level was in a remote forest area. Industrialized areas had
    high levels but so too did some agricultural regions. Higher levels
    were generally found in the western part of the study region,
    suggesting that some PCB fallout may have originated from further
    afield and be dependent on the prevailing winds. Seasonal variations
    in fallout correlated well with precipitation. Lower levels of PCB
    precipitation were found in Iceland by Bengtson & Sdergren (1974).
    The highest level was found in Northern Iceland at 1050 ng/m2 per
    month and, like other sites sampled, showed a seasonal trend with
    highest levels in the summer.

    Harvey & Steinhauer (1974) measured PCBs in the atmosphere over the
    western North Atlantic. They found that concentrations decreased
    exponentially with distance from land and concluded that wind
    transport is the major method of transport over the oceans. They also
    suggested that PCBs are transported primarily in the vapour phase. Dry deposition

    Atmospheric input into the Great Lakes has been studied extensively,
    because the lakes, as a whole, represent the largest surface area of
    any freshwater body in the world, with the lake surface area
    comprising from 27% (Ontario) to 64% (Superior) of the total basin
    area, and ranging from 19 000 km2 (Ontario) to 82 100 km2
    (Superior). Eisenreich et al. (1981) estimated that more than 80% of
    the annual mean total input of PCBs in Lake Michigan originated from
    the atmosphere. They estimated that approximately 56% of the
    9000 kg/year of PCB input in Lake Michigan was in the form of wet
    deposition and that 30% of the 6600-8300 kg/year input in Lake
    Superior was also in this form. However, Andren (1982) calculated a
    precipitation input of 650 kg/year for Lake Michigan, again assuming
    that all PCBs were on 0.5 m airborne particles. Even assuming the
    lowest estimate for the annual input of PCBs into the lake,
    approximately 60% of the total input might be atmospheric deposition.

    Andren (1982) also measured the input of PCB into an isolated lake
    (Crystal Lake, Wisconsin), to calibrate a dry deposition model. The
    model was then applied to Lake Michigan and the author concluded that,
    assuming all particulate inputs of PCB are associated with 0.5 mm
    particles, dry deposition inputs were significantly less than wet

    Manchester-Neesvig & Andren (1989) collected and analysed air samples
    from a remote site in the Great Lakes watershed during 1984 and 1985.
    Total PCB concentrations varied from 1.82 ng/m3 in the summer to
    0.135 ng/m3 in the winter. They found that, on average, 92% of the
    PCBs detected were in the vapour phase. When these data were compared
    with data collected over the previous 7 years, no significant changes
    in PCB concentrations were found. The authors concluded that, on the
    basis of the short residence time and the relatively constant annual
    average levels of PCBs, repeated cycling between earth and atmosphere
    takes place.

    Murphy (1984) reviewing data from the Great Lakes region on the
    relative distribution of airborne PCBs between particulate matter and
    vapour, concluded that they are transported predominantly in vapour.
    He stated that there was reasonable evidence to suggest that the
    atmosphere is the major source of the PCBs found in Lakes Michigan,
    Superior, and Huron, Siskiwit Lake on Isle Royale, and probably in the
    upper Great Lakes too.

    Using liquid-coated collecting plates in near-shore areas of Lakes
    Huron and Michigan, close to urban centres, more PCBs were found on
    the upper plates suggesting that much of the dry deposit of PCBs was
    associated with large particles (20 m). This sampling technique also
    indicated that, for the areas studied, dry deposition inputs were
    higher than wet inputs (Murphy, 1984).

    Duinker & Bouchertall (1989) analysed filtered air, particulates, and
    rain, in the city of Kiel, Federal Republic of Germany for 14
    different PCB congeners. They found that congeners with a low degree
    of chlorination were dominant in filtered air, whereas, congeners with
    a high degree of chlorination dominated in aerosols and rainfall. The
    vapour phase represented up to 99% of the more volatile congeners
    (i.e., those with a lower degree of chlorination). The particulates
    were found to carry relatively more of the less volatile congeners.
    Particle scavenging was the dominant source of PCBs in rain water
    despite the small contribution of particulate PCBs to the overall
    atmospheric concentration of PCBs (only 1 or 2%).

    In a study by Sdergren (1973), most of the PCB deposited on a south
    Swedish lake was in the form of dry deposit, with 11% as particulate
    matter in the precipitation and 2% from precipitation water. McClure
    (1976) stated that, on the basis of flux measurements and model
    calculations, most of the PCB fallout is in the form of dry deposition
    and that most of the dry deposition of aerosol PCB introduced into the
    troposphere falls within 100 km of its source.  Precipitation deposition

    Precipitation scavenging of chlorinated hydrocarbons in the atmosphere
    is complex. Scavenging of particles by cloud droplets and by rain
    drops in, and below, clouds, and the scavenging of the vapour phase by
    rain occurs (Murphy 1984). Thus chlorinated hydrocarbons are
    concentrated in precipitation rather than in the atmosphere, resulting
    in rainfall levels of many ng/litre. Swain (1978) and Strachan &
    Huneault (1979) measured levels in rainfall ranging between 0 (not
    detectable) and 230 ng/litre in the Great lakes area.

    Murphy (1984) pointed out that variables, such as the amount of
    particulate material and PCBs in the atmosphere, the type of rain, and
    the rate of rainfall, will affect the precision of precipitation

    Levels of PCBs in the rainfall throughout Canada during 1984 were
    monitored by Strachan (1988). Levels ranged from nd to 17 ng/litre, no
    geographical trends were apparent.

    4.1.2  Transport in soil

    PCBs in soil, derive from particulate deposition (often concentrated
    in urban areas), wet deposition, the use of sewage sludge as a
    fertilizer, and leaching from landfill sites.

    Significant amounts of PCBs are deposited on soil by particulate
    deposition (see previous section). Fujiwara (1975) analysed soil
    samples in Japan, and found that the main sources of PCB contamination
    of agricultural soils are the industries using PCBs. Other sources
    include treatment of soil with sewage sludge and accidental spills.
    The 15% of soil samples in Indiana (USA) that contained more than
    50 mg/kg had been treated with PCB-contaminated dried sludge (Bergh &
    People, 1977).

    Tucker et al. (1975a) found that, during a 4-month period following
    the addition of Aroclor 1016 to soil, the PCBs were not readily
    leached by percolating water and that only the lower chlorinated
    isomers were leached. The ease of leaching from different soils was in
    the order sandy loam > silty loam > silty clay loam.

    The behaviour of 14C-labelled PCB in flooded soils was studied by
    Ogiso et al. (1976). The amounts of PCB volatilized occurred in the
    following order: water > subsoil > soil. The addition of compost
    powder to soil reduced the amount that volatilized.

    Haque et al. (1974) studied the adsorption of Aroclor 1254 on various
    soil particle types in an aqueous solution of 56 g PCB/litre.
    Delmonte sand and silica gel did not adsorb any PCB. Woodburn soil
    adsorbed the highest amount followed by illite, montmorillonite, and
    kaolinite clays, in decreasing order. The high adsorptive capacity of
    Woodburn soil was attributed to the presence of organic matter and
    lipophilic or hydrophobic materials. Moza et al. (1976a) found that, 2
    years after the application of 14C-labelled dichlorobiphenyl to a
    loamy sand soil at 1 mg/kg, most of the detectable PCB was in the top
    10 cm of the soil and only 0.2% had reached a depth of 40 cm. In
    another study, Suzuki et al. (1977) found that Aroclors 1242 and 1254
    did not move upwards through uncontaminated sand deposited over
    contaminated soil. The leaching of water from soil may lead to a
    downward movement of PCBs, depending on the soil type and clay content
    (Pal et al., 1980).

    A large spill of  Askarel (containing 70% Aroclor 1254 and 30% tri-
    and tetrachlorobenzenes) occurred at a transformer-manufacturing
    facility in Canada, in 1976. Condie silt from near the site of the
    spill was studied with respect to the sorption partition coefficients
    and the transport retardation factors. The sorption partition
    coefficient values for 2,5,2',5'-tetrachloro-, 2,4,5,2',5'-penta-
    chloro-, and 2,4,5,2',4',5'-hexachlorobiphenyl were 5000, 9400, and
    26 000, respectively. The mean transport retardation factors for these
    3 congeners were 2.7 E + 04, 5.0 E + 04, and 1.4 E + 05, respectively.
    This implies that dissolved PCBs will move only very slowly through
    unfractured Condie silt (Anderson & Pankow, 1986).

    4.1.3  Transport in water

    PCBs enter water mainly from discharge points of industrial and urban
    wastes into rivers, lakes, and coastal waters. In static water, PCBs
    are more concentrated in the surface micro-layer than in subsurface
    samples (Bidleman & Olney, 1974). This is probably due to deposition
    from the air rather than redistribution in the water. On account of
    their low water solubility and high specific activity, it is expected
    that most of the PCBs discharged will be adsorbed by sediment at the
    bottom of rivers or lakes and transport will be mainly via waterborne
    particles (Nisbet & Sarofim, 1972). The bulk of the PCBs will sink to
    the bottom sediments. The sinking rate of PCBs from the surface to
    deeper layers in the open ocean is relatively slower in tropical
    waters than in high-latitude waters (Tanabe, 1985).

    Oloffs et al. (1973) added 0.1 mg Aroclor 1260/litre to water samples
    in the presence of sediment. After 6 weeks, all of the PCBs had been
    adsorbed by the sediment, none being given off to the atmosphere. The
    degree of PCBs sorption is inversely related to the size of the
    particles (Haque et al., 1974) and the solubility of PCBs in water
    (Haque & Schmedding, 1975). Smaller particles have a relatively larger
    surface area and so adsorb more PCBs (Steen et al., 1978). Nau-Ritter
    et al. (1982) found the adsorption and retention of PCBs to be
    directly related to the particle organic content. A significant
    correlation was found by Larsen et al. (1985) between PCB levels and
    total organic carbon in the deepwater sediments of the Gulf of Maine,
    PCBs were concentrated on finer grain particles. Organic carbon and,
    therefore, the PCB concentration were also correlated with depth.
    Wildish et al. (1980) found that estuarine sediments, especially those
    containing higher levels of organic matter, readily adsorbed Aroclor
    1254. The PCBs were found to be tightly bound to the sediment with
    virtually no desorption. Horzempa & Di Toro (1983) found that the
    adsorption of hexachlorobiphenyl was correlated with both sediment
    surface area and organic content. Adsorption was found to be
    significantly greater at 40C than at 1C. Hexachlorobiphenyl is
    strongly adsorbed on sediment and weakly desorbed. There is no simple
    reversible reaction.

    Fisher et al. (1983) found that the rate of release of PCBs from
    contaminated sediment was a function of sediment PCB concentration,
    chlorine substitution pattern, and degree of chlorination. In the
    absence of disturbance, even very low deposition rates of new sediment
    will quickly remove PCB-contaminated sediments from diffusional
    communication with overlying water. Little change was found (Nimmo et
    al., 1971a) in the PCB concentration in sediment at a point downstream
    of a contamination source over a period of 9 months. The very small
    amounts of PCBs leached from sediment into overlying water may be
    taken up by organisms.

    Hom et al. (1974) stated that the annual inputs of PCBs into the
    southern California bight from waste water and from surface runoff in
    1970-71 were estimated to be 10 and 0.25 tonnes, respectively.

    Sewage treatment appears to remove PCBs from waste water,
    concentrating them in the sludge. However, often, the sludge is then
    discharged into open water (Ahling & Jensen, 1970). Holden (1970)
    found an average of 3 mg PCBs/kg in wet sewage sludge dumped in the
    Clyde estuary, in the United Kingdom, and calculated that this would
    be equivalent to approximately one tonne per year. A similar annual
    discharge of PCBs in the sludge on the Californian coast was
    calculated by Schmidt et al., (1971).

    Dredging of inland rivers and harbours may lead to a significant
    transfer of PCBs from contaminated sediments, especially when dumped
    at sea (Nisbet & Sarofim, 1972). Rice & White (1987) found that there
    was an increase in water concentrations of PCBs immediately following
    the dredging of sediment in the Shiawassee River, Michigan. The
    availability of PCBs for clams and fish, as measured by an increase in
    uptake, was found for up to 6 months following dredging.

    4.1.4  Transport between media

    In a model ecosystem, Sdergren & Larsson (1982) found that the
    presence of bottom-living organisms, such as  Chironomus and
     Tubifex, resulted not only in the uptake of PCBs from the sediment
    but also in the release of PCBs into the water and to the surface
    microlayer, compared with a system without organisms. PCBs were
    transported to the air via jet drops from bursting bubbles in the
    surface microlayer.

    A similar pattern was found using large outdoor artificial ponds
    (Larsson, 1985a). Following the addition of Clophen A50 to sediment,
    the transport of PCBs from sediment to water followed a seasonal
    cycle, with higher levels in the summer than in the winter. The
    processes that transfer PCBs across the sediment/water interface
    (bioturbation, desorption, and gas convection) are positively related

    to temperature. Transfer from water to air was probably dominated by
    volatilization with maximum concentrations of PCBs in air at the
    highest water concentrations, lower chlorinated biphenyls achieving
    the highest concentrations in air. The majority of the airborne phase
    was presumed to be in the gaseous phase as it passed through particle
    filters. In the same ponds, Larsson & Okla (1987) measured the rate at
    which PCBs volatilized from water to air. PCB compounds volatilized at
    a rate of 0.9 to 9.6 ng/m2 per h, the rate increasing with the
    temperature of the water and the concentration of PCBs. The transport
    rate during the day exceeded the rate at night and was positively
    correlated with the air temperature (Okla & Larsson, 1987).

    Larsson (1985b) added Clophen A50 to the sediment in a model ecosystem
    comprising sediment, water, benthic macroinvertebrates, and fish. PCBs
    were detected in the water. The transport of PCBs from the water to
    air included at least 2 routes, volatilization and jet drop transport.
    Both routes were of the same magnitude (0.2-1.0 g/week). However,
    though the PCBs transported by volatilization consisted of lower
    chlorinated isomers, those transported by jet drops were identical to
    those in the sediment and water.

    In an earlier study, Larsson (1984) measured the uptake of PCBs from
    sediment by chironomid midge larvae and the concentrations of PCBs
    from larva to adult. In the field, chironomid larvae contained
    114 g/kg fresh weight at a sediment concentration of 39 g/kg wet
    weight. Different sediments affected the amount of PCBs available to
    the organisms. Adult chironomids sampled near a sewage plant contained
    251 g/kg fresh weight. The chironomid larval population was estimated
    to be 9900 per m2 and the authors calculated that these would move
    20 g PCB/m2 per year into the terrestrial compartment of the

    A model, based on the fugacity concept, was described and illustrated
    by applying it to the time-varying fate of PCBs in Lake Ontario over
    the period 1940-2000. Expressions are included for a great number of
    variables, such as loadings and the partitioning of the contaminant
    between the phases of air, aerosols, water, suspended and bottom
    sediments, various trophic levels of aquatic organisms, and gull eggs.
    Also included are expressions for transformation rates, and transport
    rates for diffusion between water and sediment, and water and air wet
    and dry atmospheric deposition, sediment deposition, burial, and
    resuspension, and water and the inflow and outflow of suspended
    matter. The results obtained by numerical integration and by assuming
    reasonable loading and air concentrations were in accordance with
    data. It was shown that PCBs cycle appreciably between the atmosphere
    and water by wet and dry deposition and volatilization, and between
    water and sediment by deposition, resuspension, and diffusion.
    Biomonitors were shown to be particularly valuable indicators of
    contamination levels in the ecosystem (MacKay, 1989).

    4.2  Biotransformation

    4.2.1  Biodegradation

    Nissen (1973) did not find any alteration in Aroclor 1254 after a
    9-week incubation period in soil. Iwata et al. (1973) added Aroclor
    1254 to various soil types. They did not find any change after one
    year in soils containing high amounts of organic matter (10.8-19.5%).

    Biotransformation had occurred, causing the disappearance of the lower
    chlorinated biphenyls, in soils with a low organic matter content
    (0.1-3.3%), as diverse as loamy sand and clay. The authors concluded
    that, after one year, the material remaining in loamy sand (0.1%
    organic matter) consisted of mainly penta- and hexachlorobiphenyl
    isomers.  Bacteria

    The biodegradation of PCB isomers, which is possible with some aerobic
    bacteria, depends on the degree of chlorination and the position of
    chlorine substitution. Degradation decreases with increasing
    chlorination. Dechlorination of PCBs occurs in anaerobic sediments.
    Here bacterial activity is preferentially targeted towards PCB
    congeners with higher levels of chlorination. Products of
    dechlorination are, therefore, more readily degraded by aerobic

    Early experiments were carried out to study the biodegradation of PCBs
    using activated sludge inocula; some degradation was found (Baxter et
    al., 1975). However, the presence of PCBs in sewage sludge shows that
    they are not all readily transformed by microorganisms. Fries (1972)
    analysed silage containing PCBs (Aroclor 1254) that had undergone
    normal fermentation. The gas chromatogram of the standard was
    identical to that of the silage sample. The authors suggested that, if
    anaerobic degradation had taken place, it would have been unlikely to
    have been uniform for all components. They stated, however, that this
    test may not have been a good indication of possible anaerobic
    degradation because DDT showed much less degradation, under the same
    conditions, compared with other degradation test systems.

    Lunt & Evans (1970) postulated a metabolic pathway, used by
    microorganisms, for biphenyl oxidation, which was later confirmed by
    the findings of Gibson et al. (1973) using a bacterium isolated from a
    polluted stream. Lunt & Evans (1970) found that a Gram-negative
    bacterium oxidized biphenyl to phenylpyruvic acid with the
    intermediary formation of 2,3-dihydroxybiphenyl and
    alpha-hydroxy--phenylmuconic semialdehyde. Catelani et al. (1971)

    found that the metabolism of biphenyl by  Pseudomonas putida was
    different, in that, though the intermediate products were the same,
    benzoic acid was isolated, not phenylpyruvic acid. Ahmed & Focht
    (1973a) isolated 2 species of  Achromobacter from sewage effluent
    using biphenyl and  p-chlorobiphenyl as the sole carbon source. They
    found that both sources were rapidly degraded, biphenyl being oxidized
    to benzoic acid and both mono and dichlorinated biphenyls to
     p-chlorobenzoic acid. In a second study, Ahmed & Focht (1973b)
    investigated the biodegradation of other isomers of PCBs, with 2-5
    chlorine atoms. The extent of oxidation seemed to be somewhat
    dependent on the presence of unsubstituted biphenyl rings. Because of
    the absence of chloride in all the supernatants, they concluded that
    the bacterium was unable to dechlorinate the PCBs. The fact that
    increasing chlorine substitution rendered the molecule more resistant
    to microbial attack was used to support this argument. However, Kaiser
    & Wong (1974), studying the degradation of Aroclor 1242 by a bacterial
    culture, isolated from lake water, showed that the PCBs were degraded
    into several metabolites (aliphatic and aromatic hydrocarbons), none
    of which contained chlorine. Dechlorination had already taken place at
    an early stage of metabolism.

    Wong & Kaiser (1975) found that lake water bacteria could use both
    Aroclor 1221 and 1242, but not 1254, as a sole carbon source for
    growth, but that only 1% of the bacterial culture had this ability.
    The authors then followed the degradation of Aroclor 1221. After one
    month, the mixture had been totally degraded to several compounds of
    low relative molecular mass. Unchlorinated biphenyls were degraded
    faster than chlorinated forms.

    Tucker et al. (1975b) observed the degradation rates of Aroclors 1221,
    1016, 1242, and 1254, and MCS 1043 (a non-commercial mixture). They
    found a clear relationship between the level of chlorination and the
    relative degradability, when degradation rate was plotted against
    percentage chlorine by weight. Volatilization rates fell within the
    95% confidence limits of overall disappearance rates and so could be
    ruled out. Analysis of the Aroclors, following exposure to the
    activated sludge, revealed a redistribution of the dominant PCBs. For
    example, the chromatograms for Aroclor 1221 and 1242 were very similar
    showing that the lower chlorinated biphenyls were more rapidly
    degraded. Furthermore, since Aroclor 1221 was found to be rapidly
    degraded, a closer study was performed that showed that most of the
    degradation occurred within 24 h.

    The degradation of polychlorinated biphenyls by either  Nocardia spp.
    or  Pseudomonas spp. was studied by Baxter et al. (1975). They found
    that, under experimental conditions, many of the lower chlorinated
    biphenyls (<3 chlorine atoms/molecule) were degraded very readily
    and some biphenyls containing as many as 6 chlorine atoms could be
    degraded, if the conditions were suitable. When PCB mixtures Aroclor
    1016 and 1242 were used, a different pattern of degradation was
    observed with an enhanced ability of the microorganisms to degrade.
    For example, 4,4'-dichlorobiphenyl degraded to 50% in about 2 days,
    when presented to  Nocardia spp. as a component of Aroclor 1242, but
    it was virtually unaffected after 12 days exposure as the pure isomer.
    The authors suggested that mutual solubilization might play some part.

    Sayler et al. (1977) found that an estuarine  Pseudomonas sp. was
    able to degrade both mixtures of PCBs (Aroclor 1254) and pure isomers
    of hexachlorobiphenyl. Degradation was dependent on incubation time
    and the purity and degree of chlorination of the biphenyl. Appreciable
    degradation occurred at all substrate concentrations of the Aroclor
    (10, 100, and 1000 g/litre) within 22 days. Although, over this
    22-day period, only 9% had been degraded at the lowest concentration
    compared with 30-40% for the other concentrations, after 60 days, this
    was reversed with 84% being degraded at 10 g/litre, 70% at
    100 g/litre, and 63% at 1000 g/litre. When compared with the pure
    isomer, degradation of the Aroclor mixture proceeded at a slower rate.
    Even though average chlorination was less, the authors speculated that
    this could be owing to the substitution positions of the chlorines.
    Chromatographic tracings showed that degradation of the lower
    chlorinated components of the Aroclor occurs before degradation of the
    more highly chlorinated biphenyls.

    Furukawa et al. (1978a,b) examined 31 PCB isomers (mono to
    pentachlorobiphenyl) for biodegradability by 2 bacterial species,
     Alcaligenes and  Acinetobacter. They found the following
    relationship between chlorine substitution and biodegradability.

    i.    Degradation decreased as chlorine substitution increased.

    ii.   Isomers containing two chlorines at the  ortho position of
          either a single ring or on both rings showed very poor

    iii.  Isomers, in which all the chlorines were on one ring, were
          generally degraded faster.

    iv.   Molecules with non-chlorinated rings or rings with few chlorines
          underwent preferential ring fission.

    v.    The 4'-chloro-substituted PCBs formed and accumulated a yellow
          intermediate during degradation.

    vi.   Only with respect to 2,4,6-trichlorobiphenyl was there a
          significant difference in ability to degrade between the 2
          bacteria. This compound was mostly metabolized within 1 h by
           Acinetobacter, but was degraded very slowly by  Alcaligenes.

    It was demonstrated by Carey & Harvey (1978) that mixed cultures of
    marine bacteria were capable of metabolizing both pure isomers (tri-
    and tetrachlorobiphenyl) and mixtures (Aroclor 1254). They isolated
    and partially characterized an acid lactone metabolite. They did not
    find any change in the chromatogram trace for the Aroclor but
    suggested that this might be related to the insensitivity of the
    method, since even if each of the isomers in the mixture had been
    metabolized to the same extent as pure isomers, this would still not
    have been detectable on the trace. The authors also found that no
    metabolism occurred when a chlorobiphenyl isomer in an anaerobic
    marine mud was incubated for 6 weeks. Degradation of Aroclor 1242 by
    mixed microbial cultures, isolated from soil and river water samples,
    was demonstrated by Clark et al. (1979). The predominant organisms in
    the cultures were  Alcaligenes odorans, Alcaligenes denitrificans,
    and an unidentified bacterium. The lower chlorinated isomers were not
    only degraded at a faster rate but were also more completely utilized
    by the bacteria. In general, the rate of degradation was much faster
    than in previous studies. Co-metabolism in the presence of sodium
    acetate was studied; greatly enhanced degradation was found for the
    more highly chlorinated isomers. Liu (1980) found that sodium
    ligninsulfonate also greatly enhanced the biodegradation of commercial
    PCB mixtures.

    The same author found that a  Pseudomonas sp. could oxidize Aroclors
    1221, 1016, 1242, and 1254, at a rapid rate. A kinetic study using
    resting cells revealed that Aroclor 1221 was degraded much faster
    (980 g/h per mg cell dry weight) than Aroclor 1254 (43 g/h per mg
    cell dry weight). The degradation of the higher chlorinated PCB
    (Aroclor 1254) could be enhanced by the addition of Aroclor 1221. Liu
    (1981) observed that the oxidation of Aroclor 1221 by the bacteria was
    10 times faster than with sewage. Two possible explanations for this
    difference were that the sewage contained toxic chemicals that
    inhibited the bacteria, but this was found not to be the case, or, the
    bacteria preferred Aroclor 1221 to the other substrates. This second
    explanation is a possibility, for glucose, a substrate used readily by
    most bacteria was poorly oxidized by this bacterium.  Pseudomonas
    oxidized Aroclor 1221 readily between 15 and 35C, the rate increasing
    with temperature. Reducing the temperature to 4 and 10C drastically
    retarded, but did not halt, degradation. Adjusting the concentrations
    of phosphorus and nitrogen from 2 mg to 20 mg/litre (the lower
    concentration being that found normally in sewage) did not alter the

    rate of degradation by  Pseudomonas spp. in raw sewage. But
    increasing nitrogen and phosphorus gave more reproducible results,
    suggesting that the compounds are on the border of limiting
    degradation rates in raw sewage. The oxygen content was found not to
    affect degradation at concentrations over 1 mg/litre (oxygen levels
    are generally maintained at between 2 and 3 mg/litre in activated
    sludge reactors, under the operational conditions of sewage-treatment
    plants). Liu (1982) found that, under a limited substrate supply,
     Pseudomonas spp. degraded all 7 of the major components of Aroclor
    1221. However, with excessive amounts of nutrient, preferential
    degradation of certain components was observed. The author stated that
    one of the main factors influencing this selective biodegradability
    was the position of chlorine substitution on the biphenyl.

    4.2.2  Biodegradation; individual congeners  Bacteria

    In a study by Parsons & Sijm (1988), the co-metabolism was
    investigated of several different mono-, di- and tetrachlorobiphenyls
    in chemostat continuous cultures of a  Pseudomonas strain (JB1). They
    found that chemostat conditions favoured degradation compared with
    exposure of the  Pseudomonos in batch culture, where little or no
    degradation was recorded. Using benzoate as the carbon source, results
    varied widely, with repeat incubations showing different degrees of
    degradation of chlorobiphenyls and, sometimes, no breakdown at all. In
    cultures that did degrade the materials, the monosubstituted
    4-chlorobiphenyl was rapidly degraded. Of the disubstituted
    dibiphenyls, 3,5-dichlorobiphenyl was more readily broken down than
    2,5-dichlorobiphenyl. Changing the carbon source available to the
     Pseudomonas sp. improved the reproducibility of the results. The
    authors reviewed the literature relative to their own findings and
    concluded that repeated culture on benzoate leads to the loss of the
    ability of the  Pseudomonas sp. to degrade biphenyl by  meta
    cleavage;  ortho cleavage is retained. Coding for the  meta cleavage
    resides on plasmids, which can be lost, whereas coding for the  ortho
    cleavage is chromosomal. Growth of the  Pseudomonas sp. on a
    3-methylbenzoate substrate improved degradation of the biphenyls.
    3-Methylbenzoate can only be degraded by a  meta cleavage favouring
    retention of the plasmid. Comparison of degradation of 4
    tetrachlorobiphenyls showed the influence of the positions of the
    chlorine substitutions. The relative degradability of the 5 compounds,
    shown in Fig. 3, was: 2,3,2',3'-tetrachloro- >2,5,3',4'-tetrachloro-
    > 2,5,2',5'-tetrachloro- approx. 2,6,2',6'-tetrachloro- approx.
    3,4,3',4'-tetrachlorobiphenyl. The authors stated, from the
    literature, that the first reaction in the degradation of
    chlorobiphenyls is, in most cases, 2,3-dioxygenation, eventually
    leading to the formation of chlorobenzoates. Chlorines in the  ortho
    and  meta positions will, therefore, offer steric hindrance to this

    The low degradation rate of 3,4,3',4'-tetrachlorobiphenyl is not
    explained by this mechanism, since it has 2 adjacent unoccupied 2,3
    positions, but is more likely explained by its toxicity. Steric
    influence on enzyme binding is offered as an explanation in this case.
    Similarly, Furukawa et al. (1978a) did not find any degradation of
    this compound in initial studies, though they did find degradation to
    a dichlorobenzoic acid by  Acinetobacter in a later study (Furukawa
    et al., 1978b; Rogers, undated(a)).

    FIGURE 3

    Brown et al. (1987a,b) examined patterns of PCB congeners remaining in
    sediments after spills of commercial mixtures of Aroclor. Sediment
    from 5 different sites was examined. Shifts in gas chromatographic
    peak distribution were indicative of dechlorination of congeners by
    anaerobic bacteria in the sediment. Analysis of sediment from
    different depths indicated less difference from the original traces in
    superficial layers and the greatest shift in deeper layers of the
    sediment cores. They concluded that dechlorination had taken place and
    deduced several different processes involved by comparison between
    sites. Six of these processes have been characterized in detail, each
    presumed to be mediated by different populations of anaerobic
    bacteria, with different selectivity for different congeners in the
    PCB mixture. The point of most interest was that congeners with high
    degrees of chlorination were selectively dechlorinated by these
    anaerobic organisms. Whilst dechlorination still leaves the mass of
    PCB intact, congeners with lower chlorination can be more readily

    degraded by aerobic bacteria. This anaerobic dechlorination,
    therefore, enables further degradation to take place elsewhere and
    contributes significantly to the detoxification of the PCBs. While the
    combined  meta- para selective dechlorinating/oxidizing action of
    sediment microbes for PCB residues is likely to be detoxifying, with
    respect to dioxin-like effects, there are reservations about whether
    this action would be detoxifying in respect of other, more subtle
    toxic effects of PCBs and their degradation products, known (such as
    the potential reproductive toxicity of the hydroxylated,
     ortho-enriched PCBs from sediment microbe action) and unknown. This
    is why it is important to study not only the disappearance of PCBs,
    but also the exact nature and amounts of the degradation products
    (McKinney et al., 1990). Two broad categories of transformation have
    been observed: the first dechlorinates in the  ortho, meta, and  para
    positions and the potential for the dechlorination of biphenyls is
    related to the reduction potential of the compound, the second
    dechlorinates only in the  meta and  para positions, and the
    reactivities of the congeners relate to the molecular shape. The
    second category suggested to the authors an active site on a
    dechlorinating agent that would be roughly conical with a reducing or
    hydrogenating site at the apex. In this schema,  para-substituted
    molecules could enter the site directly, enough rotation of the
    molecule would be possible for the accommodation of  meta, but not
     ortho, substitution. Quensen et al. (1988) demonstrated this
    dechlorinating capacity of anaerobic bacteria from Hudson River
    sediments in the laboratory. Dechlorination occurred primarily from
    the  meta and  para positions;  ortho-substituted congeners
    accumulated selectively. The fastest rate of dechlorination occurred
    at the highest exposure used (700 mg Aroclor 1242/kg); 53% of the
    total chlorine was removed over a 16-week incubation period. During
    incubation, the proportion of mono- and dichlorobiphenyls increased
    from 9 to 88%. The authors believed that a sequential anaerobic to
    aerobic system could be devised for the biological degradation of
    PCBs.  Fungi

    Wallnofer et al. (1973) incubated a soil fungus  Rhizopus japonicus
    in a medium containing 3H-labelled 4-chlorobiphenyl or
    4,4'-dichlorobiphenyl. After incubation for 1 week, the fungal
    mycelium was filtered out. Scans of TLC plates indicated a
    hydroxybiphenyl derivative present in the filtrate of both cultures.
    To further identify the metabolite, larger amounts of unlabelled
    4-chlorobiphenyl were added to a similar culture. The NMR and mass
    spectra were identical to a synthetic sample of 4- chloro-4'-hydroxy-
    biphenyl; mixed melting point determination showed no depression.
    Further positive identification of the product was not possible,
    because of limited material, but the experiment indicates the
    probability of degradation of biphenyl to a hydroxy derivative by a

    4.2.3  Photodegradation

    Several authors have reported that simple chlorinated biphenyls, as
    well as complex commercial PCB mixtures, undergo photoreduction in
    organic solvents (Safe & Hutzinger, 1971; Hustert & Korte, 1972; Ruzo
    et al., 1972, 1974, 1975; Sawai & Sawai, 1973; Koshioka et al., 1987)
    and aqueous systems (Crosby & Moilanen, 1973; Bunce, 1978) in the
    laboratory. Herring et al. (1972) found that PCBs degraded faster in
    hexane solution than in aqueous solution and slower in benzene

    Bunce et al. (1978) posed the question of the environmental
    significance of the photodegradation of PCBs and tried to estimate the
    likely degree of photolysis under real environmental conditions,
    rather than in solution in organic solvents at high concentrations.
    The current best estimate suggests that significant amounts,
    particularly of higher chlorinated PCB congeners, might be degraded in
    water by the action of sunlight.

    4.2.4  Bioaccumulation, distribution in organisms, and elimination

    Polychlorinated biphenyls accumulate in almost all organisms, because
    of their high lipid solubility and slow rate of metabolism and
    elimination. They accumulate preferentially in fat-rich tissues.

    Bioconcentration factors (BCFs) should be interpreted with caution,
    since they are simple ratios. The exposure concentration, therefore,
    makes a marked difference to the BCF obtained; very low exposure
    concentrations are likely to lead to high BCFs, since all the PCBs are
    absorbed, whilst high exposure concentration will tend to minimize the

    Experimental data on the bioconcentration of PCB mixtures and pure
    chlorinated biphenyls are presented in Table 9 for microorganisms,
    Table 10 for aquatic organisms, and Table 11 for plants, birds, and
    mammals.  Microorganisms

    Uptake of both pure chlorinated biphenyl isomers and commercial PCB
    mixtures by microorganisms is rapid, and high bioconcentration factors
    are achieved. While there is a suggestion in studies on some species
    that PCB congeners with higher levels of chlorination are taken up
    preferentially, in the majority of studies, all PCBs appear to be
    taken up equally. Uptake is true absorption; adsorption onto the
    surface of the organisms represents little of the uptake. Since
    resistant forms of microorganisms take up less PCBs than sensitive
    forms and dead cells accumulate more PCBs than live ones, there is
    some capacity to exclude the compounds.

    Harding & Phillips (1978b) studied the uptake of 14C-labelled
    2,4,5,2',5'-pentachlorobiphenyl, at concentrations of 0.31 or
    9.86 g/litre water, by 11 marine phytoplankton species including:
    diatoms, green algae, chrysophytes, haptophytes, and dinoflagellates.
    The cell density of each culture was maintained at 106-109 cells/
    litre. Equilibrium between water and cell concentrations of biphenyl
    was reached very rapidly after 0.5-2 h; small motile forms reached
    equilibrium within 1 h and large centric diatoms after approximately
    2 h. Exposure concentration and cell density, within the range given
    above, had little effect on the time-course of uptake. Substantial
    interspecies differences in adsorptive capacity were shown by
    differences in the Freundlich adsorption constant (log K). A large
    centric diatom,  Coscinodiscus sp., had the highest log K.  Nitzschia
    longissima, a penate diatom that has been shown to be resistant to
    PCBs (Harding & Phillips, 1978a), had the lowest log K value. The
    flagellates, with the exception of  Monochrysis lutheri, which has
    been shown to be very sensitive to the effects of PCBs, had much lower
    log K values than diatoms. Concentration factors, calculated from the
    Freundlich adsorption isotherms, ranged between 12 300 and 2 410 000.

    Biggs et al. (1980) exposed mixed species of estuarine phytoplankton
    (numerically dominated by the diatom  Skeletonema costatum) to
    14C-labelled PCB (approximately 54% chlorine by weight) at
    concentrations of 5.8 or 11.6 g/litre. At a particle concentration of
    25 mg/litre, 19-22% of the labelled-PCB was sorbed on the particles
    after a 1-h exposure, with 70-72% in the water. At 4 times the above
    particle concentration, 66-69% was sorbed on particles and only 22-23%
    was retained in the water. Doubling the amount of 14C-PCB doubled the
    mean amount of labelled-PCB in both the particles and the water. The
    authors calculated an index of sorption (the ratio of 14C-PCB sorbed
    on particles to that in an equal volume of water) at an average of
    2  1  104. The authors suggested that the higher uptake (88%) of
    PCBs found by Sdergren (1971) was probably the result of an
    unnaturally high cell concentration. Phytoplankton sampled in the
    surface waters of Long Island Sound, USA, varied seasonally in
    concentration from about 0.5 to 30 mg/litre.

    Lederman & Rhee (1982) calculated bioconcentration factors for 3
    species of Great Lakes planktonic algae (Table 9). In the case of
     Fragilaria crotonensis, the uptake of hexachlorobiphenyl into the
    frustule (the siliceous wall of the diatom) was investigated. The
    bioconcentration factors for frustules were lower by an order of
    magnitude than the factors for live and dead cells. It appears,
    therefore, that adsorption on the cell surface contributes only a
    little to the bioaccumulation of hexachlorobiphenyl.

        Table 9.  Bioaccumulation of PCBs: Microorganisms

    Organism          Biomass       Temperature   PCB type       Duration     Exposure        Bioconcentration   Reference
                      (cells/ml)       (C)                                   (g/litre)         factora

    Green alga         2  106      20-25         TeCB              1 h       10                     3200        Urey et al. (1976)
    Chlorella          2  106      20-25         HeCB              1 h       10                     7000        Urey et al. (1976)
    pyrenoidosa        2  106      20-25         OcCB              1 h       10                     1600        Urey et al. (1976)
                       2  106      20-25         DeCB              1 h       10                     5200        Urey et al. (1976)

    Algae              3.2  105                  HeCB             19 h        1                 117 000b        Lederman & Rhee
    Fragilaria         1.6  105                  HeCB             19 h        1                 313 000b        (1982)
    crotonensis                                                                                                  Lederman & Rhee

    Algae              3.4  105                  HeCB              6 h        1                 619 000b        Lederman & Rhee
    Ankistrodesmus     1.7  105                  HeCB              6 h        1                 959 000b        (1982)
    falcatus           8.5  104                  HeCB              6 h        1               1 207 000b        Lederman & Rhee
                                                                                                                 Lederman & Rhee

    Algae              1.1  106                  HeCB              6 h        1                 129 000b        Lederman & Rhee
    Mycrocystis sp.    5.5  105                  HeCB              6 h        1                 170 000b        (1982)
                       2.8  105                  HeCB              6 h        1                 264 000b        Lederman & Rhee
                                                                                                                 Lederman & Rhee

    Table 9.  (cont'd).

    Organism          Biomass       Temperature   PCB type       Duration     Exposure        Bioconcentration   Reference
                      (cells/ml)       (C)                                   (g/litre)          factora

    Fungus                          22-25         Aroclor 1254     24 h        0.007 mg/kg        1327b,c        Pinkney et al. (1985)
    Fusarium                        22-25         Aroclor 1254     48 h        0.007 mg/kg        1144b,c        Pinkney et al. (1985)

    a  Concentration of PCBs in organism/concentration of PCBs in medium or food; bioconcentration factors calculated on a wet weight
       basis unless otherwise stated.
    b  Calculated on a dry weight basis.
    c  Radioactive isotope used to calculate bioconcentration factor.

    Table 10.  Bioaccumulation of PCBs: Aquatic organisms

    Organism            Stat/     Organb   Temperature    PCB Type      Duration   Exposure      Bioconcentration    Reference
                        flowa                 (C)                                 (g/litre)        factorc

    American oyster     flow        WB                  Aroclor 1016      96 h     0.6                   6666        Hansen et al. (1974b)
    Crassostrea                     WB                  Aroclor 1254      56 d     0.01               165 000        Parrish (1973)
    virginica                       WB                  Aroclor 1254     392 d     0.01                89 000        Parrish (1973)

    Polychaete          stat        WB                  Aroclor 1254       5 d     1.1                    236        Courtney & Langston
    Arenicola marina    stat        WB                  Aroclor 1254       5 d     1 mg/kgd              0.24        (1978)

    Polychaete          stat        WB                  Aroclor 1254       5 d     1.1                    373        Courtney & Langston
    Nereis              stat        WB                  Aroclor 1254       5 d     1 mg/kgd              0.36        (1978)

    Water flea          flow        WB     20-22        Aroclor 1254      96 h     1.1               47 000e*        Sanders & Chandler
    Daphnia magna                                                                                                    (1972)

    Amphipod (M)        statf       WB                  Aroclor 1254      24 h     0.03                  8700        Pinkney et al. (1985)
    Gammarus            statf       WB                  Aroclor 1254      24 h     195.8 mg/kg          0.118        Pinkney et al. (1985)

    Scud                flow        WB     20-22        Aroclor 1254      96 h     1.6               24 000e*        Sanders & Chandler
    Gammarus            flow        WB     20-22        Aroclor 1254      21 d     1.6                27 000e        (1972)

    Glass shrimp        flow        WB     20-22        Aroclor 1254      96 h     1.3               12 300e*        Sanders & Chandler
    Palaemonetes        flow        WB     20-22        Aroclor 1254      21 d     1.3               16 600e*        (1972)

    Table 10.  (cont'd).

    Organism            Stat/     Organb   Temperature    PCB Type      Duration   Exposure      Bioconcentration    Reference
                        flowa                 (C)                                 (g/litre)        factorc

    Brown shrimp        flow        WB                  Aroclor 1016      96 h     0.9                   4222        Hansen et al. (1974b)
    Penaeus aztecus

    Grass shrimp        flow        WB     17-28        Aroclor 1254       7 d     2.3                 11 000        Nimmo et al. (1974)
    Palaemonetes        flow        WB     17-28        Aroclor 1254      16 d     1.3                 14 000        Nimmo et al. (1974)
    pugio               flow        WB     17-28        Aroclor 1254      28 d     0.62                17 450        Nimmo et al. (1974)
                        flow        WB     17-28        Aroclor 1254      35 d     0.62                26 580        Nimmo et al. (1974)
                        flow        WB                  Aroclor 1016      96 h     0.4                   2750        Hansen et al. (1974b)

    Crayfish            flow        WB     20-22        Aroclor 1254      96 h     1.2                 1700e*        Sanders & Chandler
    Orconectes nais     flow        WB     20-22        Aroclor 1254      21 d     1.2                 5100e*        (1972)

    Stonefly            flow        WB     20-22        Aroclor 1254      96 h     2.8                 2500e*        Sanders & Chandler
    Pteronarcys         flow        WB     20-22        Aroclor 1254      21 d     2.8                 2800e*        (1972)

    Dobsonfly           flow        WB     20-22        Aroclor 1254      96 h     1.1                 4600e*        Sanders & Chandler
    Corydalus           flow        WB     20-22        Aroclor 1254      21 d     1.1                 6800e*        (1972)

    Phantom midge       flow        WB     20-22        Aroclor 1254      96 h     1.3               23 600e*        Sanders & Chandler
    Chaoboruspuncti                                                                                                  (1972)

    Mosquito larvae     flow        WB     20-22        Aroclor 1254      96 h     1.5               18 000e*        Sanders & Chandler
    Culex tarsalis                                                                                                   (1972)

    Table 10.  (cont'd).

    Organism            Stat/     Organb   Temperature    PCB Type      Duration   Exposure      Bioconcentration    Reference
                        flowa                 (C)                                 (g/litre)        factorc

    Mayfly              flow        WB     8            Clophen A50        6 d     0.526                 2940        Sdergren &
    Ephemera danica                                                                                                  Svensson (1973)

    Pinfish             flow        WB                  Aroclor 1016      96 h     0.8                   2750        Hansen et al. (1974b)
    Lagodon             flow        WB                  Aroclor 1016      28 d     1 n                 25 000        Hansen et al. (1974b)
    rhomboides          flow        WB                  Aroclor 1016      56 d     1 n                 17 000        Hansen et al. (1974b)

    Sheepshead          flowg       WB                  Aroclor 1016      33 d     1 n                 26 000        Hansen et al. (1975)
    minnow              flowg       WB                  Aroclor 1016      28 d     1 n                 54 000        Hansen et al. (1975)
    Cyprinodon          flowg       WB                  Aroclor 1016      28 d     1 n                 22 000        Hansen et al. (1975)

    Spot                flow        WB                  Aroclor 1254       7 d     1 n                   7200        Hansen et al. (1971)
    Leiostomus          flow        WB                  Aroclor 1254      14 d     1 n                 17 000        Hansen et al. (1971)
    xanthurus           flow        WB                  Aroclor 1254      28 d     1 n                 37 000        Hansen et al. (1971)
                        flow        WB                  Aroclor 1254      56 d     1 n                 27 000        Hansen et al. (1971)

    Atlantic salmon     flow        WB     10-15        Aroclor 1254      33 d     10 mg/kg              0.39        Zitko (1974)
    Salmo salar

    Table 10.  (cont'd).

    Organism            Stat/     Organb   Temperature    PCB Type      Duration   Exposure      Bioconcentration    Reference
                        flowa                 (C)                                 (g/litre)        factorc

    Coho salmon         flow        WB     17           Aroclor 1254     112 d     0.048 mg/kg           9.79        Mayer et al. (1977)
    Oncorhynchus        flow        WB     17           Aroclor 1254     112 d     4.8 mg/kg             0.79        Mayer et al. (1977)
    kisutch                         WB                  TeCB              17 d     1 mg/kg              0.144        Gruger et al. (1976)
                                    WB                  TeCB              35 d     1 mg/kg              0.139        Gruger et al. (1976)
                                    WB                  PeCB              35 d     1 mg/kg              0.162        Gruger et al. (1976)
                                    WB                  HeCB              35 d     1 mg/kg              0.151        Gruger et al. (1976)

    Channel catfish     flow        WB     26           Aroclor 1232     150 d     2.4 mg/kg            1.875        Mayer et al. (1977)
    Ictalurus           flow        WB     26           Aroclor 1232     193 d     2.4 mg/kg              1.3        Mayer et al. (1977)
    punctatus           flow        WB     26           Aroclor 1248     193 d     2.4 mg/kg             0.79        Mayer et al. (1977)
                        flow        WB     26           Aroclor 1254     193 d     2.4 mg/kg                2        Mayer et al. (1977)
                        flow        WB     26           Aroclor 1260     193 d     2.4 mg/kg             1.46        Mayer et al. (1977)
                        flow        WBh    24-26        Aroclor 1242     130 d     20 mg/kg              0.72        Hansen et al. (1976a)
                        flow        WB                  Aroclor 1248      77 d     5.8                56 370*        Mayer et al. (1977)
                        flow        WB                  Aroclor 1254      77 d     2.4                61 190*        Mayer et al. (1977)

    Table 10.  (cont'd).

    Organism            Stat/     Organb   Temperature    PCB Type      Duration   Exposure      Bioconcentration    Reference
                        flowa                 (C)                                 (g/litre)        factorc

    Fathead (M)         flow        WB     25           Aroclor 1248     250 d     3           approx. 60 000        DeFoe et al. (1978)
    minnow (M)          flow        WB     25           Aroclor 1260     250 d     2.1        approx. 160 000        DeFoe et al. (1978)
    Pimephales (F)      flow        WB     25           Aroclor 1248     250 d     3          approx. 120 000        DeFoe et al. (1978)
    promelas (F)        flow        WB     25           Aroclor 1260     250 d     2.1        approx. 270 000        DeFoe et al. (1978)

    d = Days; M = Male; F = Female; DiCB = dichlorobiphenyl; TeCB = tetrachlorobiphenyl; PeCB = pentachlorobiphenyl;
    HeCB = hexachlorobiphenyl; OcCB = octachlorobiphenyl; DeCB = decachlorobiphenyl.
    a  Stat = static conditions (water unchanged for duration of experiment); flow = flow-through conditions (PCB concentration
       in water continously maintained).
    b  WB = whole body.
    c  Bioconcentration factor = concentration of PCBs in organism/concentration of PCBs in medium or food; bioconcentration
       factors calculated on a wet weight basis unless otherwise stated. * Radioactive isotope used to calculate
       bioconcentration factor.
    d  Sediment.
    e  Calculated on a dry weight basis.
    f  Static conditions, but test solution changed at intervals.
    g  Intermittent flow-through conditions.
    h  Not including stomach.

    Sdergren (1971) maintained the unicellular freshwater green alga
     Chlorella pyrenoidosa in water (at a cell concentration of
    approximately 900 mg/litre) with added nutrient medium containing
    3.7 g Clophen A50/litre, over a period of 7 days. By the end of the
    experiment, 88% of the PCBs had been taken up by the alga. The
    remaining PCBs were detected in the water, none being found in the air
    samples taken. In another study, Urey et al. (1976) found that both
    tetrachloro- and hexachlorobiphenyl isomers, at 10 g/litre, were
    concentrated by dead  Chlorella pyrenoidosa cells by 6000 and 15 000
    times, respectively, after a 1-h exposure. These concentration factors
    are approximately twice those for living cells (Table 9). Similar
    findings have been noted with other species of algae (Biggs et al.,
    1980; Lederman & Rhee, 1982).

    The ciliate  Tetrahymena pyriformis was exposed to Aroclors 1248
    (0.01, 0.1, and 1 mg/litre) and 1260 (0.001, 0.01, 0.1, and
    1 mg/litre) for 7 days (Cooley et al., 1973). Uptake of the toxicant
    increased linearly with increasing concentration. Concentration
    factors ranged from 14.8 to 40.6 for Aroclor 1248 and from 21 to 79
    for Aroclor 1260. Approximately 15-20% of Aroclor 1248 was absorbed at
    each concentration compared with means of 37-53%, with increasing
    concentration, for Aroclor 1260. If the data from Cooley et al. (1972)
    on the uptake from Aroclor 1254 is included, it is clear that
     T. pyriformis accumulates more PCBs with increasing degree of

    Dive et al. (1976) studied the accumulation of 16 pure isomers of PCB
    and one commercial product, Pyralene 3010, by the ciliate protozoan
     Colpidium campylum, at concentrations of 0.1, 1, or 10 mg/litre for
    43 h. The amount of PCBs taken up at 0.1 mg/litre was very similar for
    each of the PCB isomers and the commercial product, ranging from 29.4
    to 49%. The percentage uptake did not change greatly for the higher
    exposures.  Plants

    Uptake of PCBs into plants from soil is positively correlated with the
    soil concentration of the PCBs. Roots accumulate more than stems and
    foliage. Bioconcentration factors are low. More lower chlorinated
    congeners of the PCBs are taken up, probably because of their greater
    mobility in the soil. Much of the uptake is adsorption on the surfaces
    of roots and there is little translocation. PCBs found in leaves have
    volatilized from the soil. Uptake on root surfaces can be reduced or
    eliminated by adding activated charcoal to the soil.

    Lawrence & Tosine (1977) found that plants took up significant amounts
    of PCBs (30-140% of the applied PCB concentration) from soil treated
    with sewage sludge. In a waste PCB spill besides a North Carolina
    highway, levels as high as 4700 mg/kg were recorded in the top 3 cm of
    soil. Seven months later, the PCB concentrations were unchanged; the
    authors believed that this was because the PCBs were bound to
    activated carbon that had been used to treat the spill (Pal et al.,

    Strek & Weber (1982) analysed statistically the data from several
    literature sources (Iwata et al., 1974; Wallnofer & Koniger, 1974;
    Wallnofer et al., 1975; Iwata & Gunther, 1976; Moza et al., 1976a,
    1979a,b; Weber & Mrozek, 1979) on PCB uptake by plants, with the
    following conclusions.

    i.      The PCB content of the plant is significantly dependent on the
            soil PCB concentration.

    ii.     There is a significant difference between plant species,
            carrots taking up more PCBs than other plants.

    iii.    There appears to be a limit of the PCB concentration in the
            soil at which no detectable PCBs are taken up by the plants.

    iv.     Roots take up more PCBs than tops.

    v.      most of the PCBs in roots may, in fact, be adsorbed on the
            surface and not actually taken up.

    vi.     There is a general trend of increasing PCB content with
            decreasing chlorination, for pure PCB congeners.

    vii.    The amount of chlorination seems to have an effect on the
            mobility of PCBs within plant parts. Since lower chlorinated
            PCBs have been reported to be more mobile in soils than highly
            chlorinated PCBs, they may be more readily transported and
            available for plant uptake.

    Larsson (1987) maintained the macroalga  Cladophora glomerata in a
    flowing-water, outdoor pool. Sediment contaminated with Clophen A50 at
    2.7 mg/kg dry weight was added and PCB residues in the alga were
    monitored. The algal concentration was 3.55 mg/kg dry weight within 3
    months. Residues had fallen a year later to 0.2 mg/kg, reflecting the
    water levels of PCBs. The authors concluded that a partitioning
    process governed the uptake of PCBs by  C. glomerata in this
    experiment, because the alga accumulated the same PCBs and the same
    proportion of PCBs that were present in the water.

        Table 11.  Bioaccumulation, of PCBs: Plants, birds, and mammals

    Organism                      Organ       PCB type        Duration           Exposure   Bioconcentration   Reference
                                                                                 (mg/kg)    factora


    Beet (Beta vulgaris)          plant top   Aroclor 1254    39 days            20         0.041c             Strek et al. (1981)

    Sorghum (Sorghum bicolor)     plant top   Aroclor 1254    39 days            20         0.003c             Strek et al. (1981)

    Peanut (Arachis hypogaea)     plant top   Aroclor 1254    78 days            20         0.024c             Strek et al. (1981)

    Corn (Zea mays)               plant top   Aroclor 1254    13 days            20         0.001c             Strek et al. (1981)

    Carrot                        root        DiCBd           112 days           0.118      2c                 Moza et al. (1976a)
                                  leaves      DiCBd           112 days           0.118      0.92c              Moza et al. (1976a)


    White pelican                 carcase     Aroclor 1254    70 days            144        14.8               Greichus et al. (1975)
    (Pelecanus erythrorhynchos)

    Table 11.  (cont'd).

    Organism                      Organ       PCB type        Duration           Exposure   Bioconcentration   Reference
                                                                                 (mg/kg)    factora

    Chicken                       fat         Aroclor 1242    28 days            100        2.83               Harris & Rose (1972)
                                  fat         Aroclor 1254    28 days            100        5.15               Harris & Rose (1972)
                                  fat         Aroclor 1260    28 days            100        4.82               Harris & Rose (1972)

    Big brown bat                 carcase     Aroclor 1254    37 days            9.4        6.6                Clark & Prouty
    (Eptesicus fuscus)                                                                                         (1977)

    Mink                          fat         Aroclor 1254    approx. 56 days    1.5        16.5               Hornshaw et al.
    (Mustela vison)               fat         Aroclor 1254    approx. 126 days   1.5        28.5               (1983)
                                                                                                               Hornshaw et al.

    a  Bioconcentration factor = concentration of PCBs in organism/concentration of PCBs in medium or food; bioconcentration
       factors calculated on a wet weight basis, unless otherwise stated.
    b  Calculated on a dry weight basis.
    c  Radioactive isotope used to calculate bioconcentration factor.
    d  DiCB = dichlorobiphenyl.

    Red mangrove  (Rhizophora mangle) seedlings were grown for 6 weeks in
    soil treated with Aroclor 1242 at concentrations of between 0.038 and
    6 mg/kg (Walsh et al., 1974). Low levels (detection limit was
    0.1 mg/kg) of the PCBs were detected in the roots at exposure
    concentrations of 3 or 6 mg/kg, during the exposure period, but no
    residues were found in the stems. Residues were detected in both the
    hypocotyls and leaves at application rates of 0.3 mg/kg or more. Leaf
    residues did not change with time, but PCB concentrations in the
    hypocotyls showed an increase. The highest mean residues of 1.5 mg/kg
    were found in the hypocotyl in the highest exposure group.

    Iwata et al. (1974) treated soil with Aroclor 1254, at a concentration
    of 100 mg/kg, and sowed carrots in the plot 7 months later. The
    carrots were harvested 3 or 4 months after seeding. The authors found
    that the lower chlorinated biphenyls were more readily taken up from
    the soil into the carrot root. Analysis of the carrot peel revealed
    approximately 97% of the PCB residue, showing that there is little
    translocation within the plant; 23 months after sowing carrots in soil
    contaminated with 100 mg/kg Aroclor 1254, dissipation from soil
    appeared to parallel the degree of chlorination (Iwata & Gunther,
    1976). Analysis of the soil revealed that the lower chlorinated
    biphenyls were slowly dissipated while the more highly chlorinated
    biphenyls appeared to be unaffected. Small amounts of PCBs were found
    in carrot foliage and the authors suggested that the PCB composition
    indicated that they came from soil dust. Suzuki et al. (1977) also
    found that lower chlorinated biphenyls were preferentially taken up by
    plants, following exposure of soybean sprouts to soil contaminated
    with Aroclor 1254 or 1242 at 100 mg/kg.

    Moza et al. (1976a) found that carrot bioconcentrated
    2,2'-dichlorobiphenyl (0.118 mg/kg soil) from soil by a factor of 2
    (Table 11). No bioaccumulation was found in sugar beet, but the soil
    residue was only 0.029 mg/kg. Carrots were grown in soil amended with
    either 14C-labelled 2,5,4'-trichlorobiphenyl at 1.28 kg/ha or
    2,4,2',4',6-pentachlorobiphenyl at 1.12 kg/ha for one season (Moza et
    al., (1979a). Only 32.5% of the trichlorobiphenyl was recovered, the
    rest being lost through volatilization. The carrots had taken up 3.1%
    of the applied 14C, representing a concentration factor of 2.8. For
    the pentachlorobiphenyl, 58.5% was recovered, 1.4% of which had been
    taken up by the carrots. Sugar beet grown in the soil the following
    year accumulated only 0.4% of the applied 14C.

    In a study by Weber & Mrozek (1979), 14C-labelled Aroclor 1254 was
    applied to Lakeland soil at a rate of 20 mg/kg. Activated carbon was
    mixed with half the pots at a rate of 3.7 t/ha (3333 mg/kg). The pots
    were seeded with either soybean or fescue. After harvesting at 16 days
    for soybean and 50 days for fescue, the amounts of labelled-PCBs,
    recovered from the plant tops, were 0.016% and 0.17% for the 2

    species, respectively. The addition of activated carbon to the soil
    reduced the uptake of 14C-PCBs, the recovery of labelled-PCBs being
    0.001% and 0% for soybean and fescue, respectively. Strek et al.
    (1981) also applied 14C-labelled Aroclor 1254, at the same rate, to
    Lakeland soil; several species of crop plants were grown in the soil
    and bioaccumulation factors, calculated (Table 11). Addition of
    activated carbon (3.7 t/ha), equivalent to 3333 mg/kg to replicate
    pots, reduced the uptake of the labelled-PCBs by 80-100%.

    When approximately 1 mg 14C-labelled Aroclor 1254/kg was applied to
    the centre leaflet of the first trifoliate leaf of 18-day-old soybean
    plants, only 6.7% was recovered from the plant after 12 days, 76% of
    which was still present in the treated leaf (Weber & Mrozek, 1979).

    Mrozek & Leidy (1981) transferred the marsh plant  Spartina
     alterniflora from the field into soil containing 1 mg Aroclor
    1254/kg (dry weight) and harvested the plants after a growth period of
    90 days. The plants were found to take up selectively the lesser
    chlorinated biphenyls. The authors stated that a further shifting of
    the chromatographic pattern of PCBs towards the lesser chlorinated
    components in aerial tissues suggested that some alteration of the PCB
    mixture occurs in the plant. Mrozek et al. (1982) also found that
     Spartina accumulates PCBs from both contaminated sand and mud-soil
    systems. The total 14C-radioactivity accumulated in plants grown in
    sand systems was higher than that in plants grown in mud. The level of
    radioactivity accumulated in the green parts of the plants was similar
    in both soil systems.

    Moza et al. (1976b) applied 76 mg/kg of 14C-labelled 2,5,4'-tri-
    chlorobiphenyl or 133 mg/kg of labelled 2,4,6,2',4'-pentachloro-
    biphenyl to the leaves of the marsh plant  Veronica beccabunga. Six
    weeks later, the total recovery from plant, water, and soil was 3.7
    and 18.3%, respectively, 86 and 95% of which was recovered from the
    plant. In an earlier study, Moza et al. (1974) applied 14C-labelled
    2,2'-dichlorobiphenyl in water or soil to 2 higher water plant species
    ( Ranunculus fluitans and  Callitriche sp.) at concentrations of
    13.7 and 14.5 mg/kg, respectively. Four weeks after application, the
    results showed that the dichlorobiphenyl was metabolized more readily
    after addition to water; the authors suggested the involvement of
    aquatic bacteria. When applied in soil, 1.2% of the dichlorobiphenyl
    was metabolized. This was contributed to the plant.

    Moza et al. (1979b) grew 3-year-old spruce trees  (Picea abies) in
    soil containing 14C-labelled PCBs at approximately 4.2 mg/litre in
    sewage sludge. When analysed 4 years later, only 0.8% (0.5% in
    needles, 0.3% in stems) of the applied radioactivity was found in the
    trees. Leaching of radioactive substances from the soil was less than
    0.1% in the first 2 years and undetectable for the remainder of the

    In another study, Fries & Marrow (1981) grew soybean  (Glycine max)
    in pots, to determine residue contamination in plant tops from
    14C-labelled 2,5,2'-trichlorobiphenyl, 2,5,2',5'-tetrachlorobiphenyl
    or 2,4,5,2',5'-pentachlorobiphenyl, applied to the surface or
    subsurface soil. Each compound was added to the soil at a rate of
    2-3 mg/kg and the plants were harvested after a period of 52 days.
    Detectable residues were only found in plants grown in surface-treated
    soil, and concentrations in the plants increased with increasing
    chlorination. Little of the labelled PCBs was lost from
    subsurface-treated soil, but 20-30% of the surface-treated PCBs were
    lost through volatilization. The authors concluded that the PCB
    residues in the plant tops were, therefore, due to foliar
    contamination from vapour rather than the uptake from the soil via the
    roots. Miyazaki et al. (1975) came to the same conclusion when they
    found no absorption or translocation of PCBs in sesame or rice seeds,
    following the application of 4 types of Kanechlor (KC300, 400, 500,
    and 600) at rates of between 0.1 and 100 mg/kg. But the rice straws
    contained PCB levels of 0.02-0.08 mg/kg, which were the same as levels
    found in plants from untreated soils.

    Beets ( Beta vulgaris L.), turnips ( Brassica rapa L.), and beans
    ( Phaseolus vulgaris L.) were grown (Sawhney & Hankin, 1984) in soil
    to which lake sediment contaminated with PCBs had been added. The
    plants were exposed to Aroclor 1248 at a concentration of 80 g/kg,
    Aroclor 1254 at 1880 g/kg, and Aroclor 1260 at 14 440 g/kg. When
    beets and turnips were grown in the soil for 6 months, the plants
    showed greater uptake in the leaves than in the roots. For example,
    beet roots contained 15, 16, and 35 g/kg of Aroclors 1248, 1254, and
    1260, respectively, while beet leaves contained 22, 94, and 52 g/kg,
    respectively. Total concentrations of the 3 Aroclors in beet roots and
    leaves and in turnip roots and leaves were 66 and 168 g/kg,
    respectively, and 66 and 99 g/kg, respectively. During a second
    growing season, turnips and beans were grown for 6 months without any
    additional PCB-contaminated sediment. Comparing PCB levels in turnips
    between the 2 growing seasons showed a decrease in Aroclor 1248 uptake
    relative to Aroclors 1254 and 1260. This was primarily because of a
    large reduction in the amount of Aroclor 1248 in the soil after 1
    year, due to degradation and volatilization. In beans, higher PCB
    levels were found in the leaves and pods than in the stems and seeds.

    Ten sludge application sites were selected within the Ontario area to
    determine background heavy metal and PCB concentrations in the soils
    and crops. Control sites (without sludge application) were adjacent to
    the sludge application sites. Grab samples of liquid sludges applied
    at each of the sites were taken for analysis. The soil samples were
    taken at a depth of 15 cm. Twenty core samples were taken at 20-m
    intervals and combined to form 1 sample. Eight of the application

    sites were cropped with corn, one with oats, and one was left without
    a crop. At the control sites, 7 were cropped with corn, 1 with oats,
    and 2 left without a crop. PCB concentrations in the sludges ranged
    from 0.13 to 1.61 mg/kg dry solids. PCB concentrations were in the
    range of 0.007-0.025 mg/kg in the soil without sludges, and in the
    range of 0.018-0.453 mg/kg air-dry weight in the soil with sludges.
    The PCB levels in the crops were close to the control values (Webber
    et al., 1983).

    Bacci & Gaggi (1985) assessed the influence of translocation on the
    concentrations of PCBs in the foliage of different plant species.
    Beans, broad beans, tomatoes, and cucumbers were grown, either in soil
    with a nominal added concentration of 500 mg/kg Fenclor 64 (similar to
    Aroclor 1260), or in clean sand, for 28 days, enclosed in a glass box
    with a constant turnover of air. The plants grown in clean sands were
    exposed to PCBs by volatilization from other pots containing PCBs,
    which were in the same growing box. The PCB peak pattern of both sand
    and roots was similar to that of Fenclor 64, whereas the peak pattern
    for foliage and air had moved towards lesser chlorinated congeners.
    The concentrations of PCBs in the roots of tomatoes grown in
    contaminated soil ranged from 105 to 168 mg/kg dry weight. But
    translocation through the plants does not seem to be very likely since
    there was no significant difference in foliage uptake of PCBs between
    plants grown in contaminated soil and plants grown in clean soil.
    Foliar uptake ranged from 13.8 to 42.6 mg PCB/kg (dry weight) for the
    different species in PCB-fortified soil and from 11.8 to 47.1 mg/kg
    for plants grown in clean soil.  Aquatic invertebrates

    Bioconcentration factors are high for PCBs taken up by aquatic
    invertebrates exposed to either pure chlorinated biphenyl isomers or
    commercial mixtures in the water. Since PCBs are strongly bound to
    sediments, this method of exposure is unrealistic. Addition of
    sediment to test tanks decreases the uptake of PCBs, particularly by
    organisms living in the upper water. However, there is clear evidence
    that PCBs can also be readily absorbed into invertebrates from both
    sediment and food. For organisms living on or in, sediment, uptake can
    take place from the sediment, via food organisms that have absorbed
    the PCBs, and from interstitial water or water immediately above the
    sediment layer. A high content of organic matter in sediment decreases
    the availability of PCBs for organisms. Uptake is rapid in most cases
    and equilibrium is often reached in hours, though it may take weeks in
    other examples. Uptake increases with increasing temperature. The
    route of uptake is often via the gills, but varies among species. Loss
    of PCBs is slow, but residues do decrease on cessation of exposure.
    PCB uptake by aquatic invertebrates is transferred to predators and
    can also be transferred to the terrestrial environment.

    (a) Uptake from water

    Vreeland (1974) exposed American oysters  (Crassostrea virginica) to
    various PCB isomers at concentrations of 5.5, 17, or 60 ng/litre
    (which is within the range found in coastal waters) for 65 days.
    Equilibrium was reached after approximately 1 month of exposure, with
    concentration factors ranging from 1200 to 48 000 for PCB isomers with
    2-6 chlorine atoms/molecule. The PCB concentration, after equilibrium
    had been reached, was directly proportional to the amount of PCBs
    added to the water. Lowe et al. (1972) found a linear pattern of
    uptake in young American oysters exposed to Aroclor 1254 at 5 g/litre
    for 24 weeks, followed by a further 32 weeks in clean water. The
    oysters already contained 17 mg/kg from a previous exposure and, by
    the end of the 24-week exposure period, had accumulated 425 mg/kg (a
    steady state was not established). By the end of the 32-week period in
    clean water, no PCB residues could be detected. In another study on
    uncontaminated young oysters, concentration factors of up to 101 000
    were achieved after a 25-week exposure to 1 g Aroclor 1254/litre.
    After 12 weeks in clean water, whole-body residues were reduced to
    0.2 mg/kg.

    Courtney & Denton (1976) fed hard clams  (Mercenaria mercenaria)
    Aroclor 1254 adsorbed on the surface of alumina particles, at 1.25 and
    12.5 g/litre, for 21 days. The maximum concentration factor was 1800
    for visceral mass, when the clams had been exposed to 1.25 g/litre
    for 18 days. The visceral mass accumulated a 1.4-5.3 times greater
    concentration of PCBs per unit time than the muscular foot. Tissue
    samples contained relatively more lower chlorinated isomers than the
    Aroclor 1254 standard and, faeces and mud samples contained more
    higher chlorinated isomers. Following exposure, clams from the lowest
    dose group showed little change in PCB content after 3 months in clean
    seawater. However, at the higher dose level, there was a significant
    reduction in the PCB levels found in the foot after 1 month, but PCB
    residues in the visceral mass remained unchanged for 6 months.

    Pink shrimp  (Penaeus duorarum) were exposed to Aroclor 1254 at a
    concentration of 2.5 g/litre, in flowing water, for 22 days (Nimmo et
    al., 1971b). Accumulation was linear for the hepatopancreas and whole
    body, but a plateau was reached after 2 days in muscle. Residues in
    the hepatopancreas reached 510 mg/kg over the exposure period,
    representing a concentration factor of 204 000; over the same period,
    50% of the shrimps died. In a separate study, the shrimps were exposed
    to 7.5 g/litre for 16 days followed by an elimination period of 5
    weeks in clean water. When calculated on the basis of the total tissue
    burden of PCBs, an 80% reduction in the hepatopancreas was found,
    concomitant with a doubling of the PCB levels in remaining tissues.
    However, when data were presented as a concentration, a linear loss
    from the hepatopancreas was seen, with the concentration in other
    tissues remaining constant. The authors calculated a half-life for
    loss of PCBs from the hepatopancreas of 17 days.

    Nimmo et al. (1975) sampled shrimp from Pensacola estuary, USA, and
    measured the relative concentration of PCBs in the various tissues.
    The hepatopancreas contained the greatest amounts (50-75%) followed by
    the ventral nerve. The authors studied the uptake of PCBs by pink
    shrimp, experimentally, using various regimes with dosed food or dosed
    water. They found the same tissue distribution in pink shrimp that had
    been exposed to 0.2 g Aroclor 1254/litre, in water, for 50 days. They
    concluded that most of the PCBs were taken up directly from the water
    in both the "wild" and laboratory situation. However, they did not
    exclude the possibility of some PCBs being taken up from food, which
    was found under some of the laboratory regimes.

    To determine whether there was a concentration below which shrimps
    would be unable to accumulate PCBs, grass shrimp  (Palaemonetes pugio)
    were exposed to flowing water concentrations of 0.04, 0.09, or
    0.62 g/litre. Whole-body residues of 0.2, 1.0, and 10 mg/kg,
    respectively, were accumulated within 3-5 weeks. Even at the lowest
    dose, shrimps accumulated more PCBs than the residues found in control
    shrimp. Concentrations in the shrimp did not reach equilibrium during
    the 5-week exposure, but the rate of accumulation decreased towards
    the end of the exposure. When transferred to clean water, the shrimps
    lost most of the PCBs within 4 weeks (Nimmo et al., 1975).

     Gammarus oceanus were exposed by Wildish & Zitko (1971) to Aroclor
    1254 concentrations of 2.5, 10, or 20 mg/litre for up to 6 h. Uptake
    increased with increasing PCB concentration. Uptake decreased to half
    of the initial rate after 4-6 h exposure at 20 mg/litre. There was
    little or no uptake by dead animals. Although uptake was related to
    branchial surface area, branchiae were not necessary sites of uptake,
    since uptake could occur at an unchanged rate following branchial
    removal. The authors did not find any change in the rate of uptake
    during the intermoult stage.

    Zhang et al. (1983) exposed  Daphnia magna to 14C-labelled
    2,2'-dichlorobiphenyl, 2,5,4'-trichlorobiphenyl, 2,4,6,2'-tetra-
    chlorobiphenyl, or 2,4,6,2',4'-pentachlorobiphenyl at 50 g/litre.
    Equilibrium was reached after 20 h for all except the pentachloro-
    biphenyl, which had not reached equilibrium within 24 h.
    Bioaccumulation factors at equilibrium ranged from 3741, for the
    dichlorobiphenyl, to 18 144, for the trichlorobiphenyl. Concentration
    factors were not significantly related to the water solubility or
    chlorine content of the biphenyl, but there was a tendency for the
    bioaccumulation factor to increase with chlorine content and
    decreasing water solubility. The authors studied the rate of
    depuration and found it to increase with increasing water temperature
    between 2 and 22C. The rate of depuration was also faster for the
    dichlorobiphenyl than for the pentachlorobiphenyl; after 48 h, the
    amount of PCBs remaining in  Daphnia was 22% and 77% (at 10-11C) for
    the 2 chlorobiphenyls, respectively.

    (b) Uptake from sediment

    Sediment was collected from the field and spiked with Phenochlor DP-5
    to achieve a final PCB concentration of 0.65 mg/kg dry weight,
    compared with 0.2 mg/kg in unspiked sediment (Elder et al., 1979).
    Worms  (Nereis diversicolor) were then added to aquaria containing
    the sediment under flowing seawater. Equilibrium was reached within
    40-60 days, by which time both groups had concentrated the PCBs by 3.5
    times. Upon transfer from spiked to unspiked sediment, the worms took
    2 months to attain body levels of PCBs comparable with those of the
    unspiked group. A half-life of approximately 27 days was calculated
    for incorporated PCBs.

    Fowler et al. (1978) exposed  Nereis diversicolor to spiked sediment
    containing 9.3 or 80 mg Phenochlor DP-5/kg (dry weight), for 120 days,
    compared with 0.11 mg PCB/kg in unspiked sediment. At the beginning of
    the study, worms in the unspiked sediment had body residues of
    0.59 mg/kg dry weight and reached a steady state at 1.2 mg/kg. Those
    exposed to spiked sediment reached a steady state after a period of
    approximately 2 months, with concentration factors ranging from 3 to
    4. The worms maintained at the highest level of PCBs all died within a
    90-day exposure period. When transferred to unspiked sediment for a
    2-month period, the worms that had taken up PCBs from the unspiked
    sediment lost PCBs exponentially. In a separate study, worms were
    exposed to PCBs in water alone at a concentration of 0.57 g/litre. A
    steady state was reached much more quickly (2 weeks) than it was in
    the presence of sediment, with a concentration factor of approximately
    800. By comparing these results with field monitoring, the authors
    calculated the relative importance of the 2 media. They stated that
    approximately 99% of the PCBs in these studies was taken up from the
    sediment. When the water overlying the spiked sediment was monitored,
    28 ng PCBs/litre was measured (not leached, but a contaminant in the
    experimental system) reducing the figure of uptake from sediment to

    In a study by Courtney & Langston (1978), 1.1 mg Aroclor 1254/kg was
    incorporated into intertidal sand. Specimens of 2 intertidal
    polychaetes  (Arenicola marina and  Nereis diversicolor) containing
    mean residues of 0.017 and 0.11 mg PCBs/kg (wet weight), respectively,
    were collected. After 5 days in the spiked sediment, they contained
    0.24 and 0.36 mg/kg, and, after a further 5 days, 0.39 and 0.49 mg/kg,
    respectively. During a 3-week post-exposure period, there was no
    significant loss of these PCB residues. The authors achieved
    comparable PCB residues in these polychaetes after exposure to
    1 g/litre water or 1 mg/kg sediment.

    McLeese et al. (1980) exposed the polychaete worm  (Nereis virens)
    and the shrimp  (Crangon septemspinosa) to sediment containing
    0.016-0.58 mg Aroclor 1254/kg (dry weight) for 32 days. Uptake was
    found to be dependent on the exposure concentration and, in the case
    of the worms, on the exposure period. The accumulation of PCBs was
    inversely related to animal size; at 32 days, concentration factors
    for worms ranged from 10.8 for 0.6-g worms to 3.8 for 4.7-g worms
    following exposure to 0.17 mg PCB/kg. Factors of 3.5 and 1.9 were
    found for shrimps weighing 0.1 and 2.9 g, respectively, after exposure
    to 0.13 mg Aroclor/kg. Shrimps were found to accumulate, on average,
    60% less PCBs than worms per unit weight. During the 26 days following
    exposure, there was not any obvious loss of PCBs from the worms.

    Rubinstein et al. (1983) collected sediments containing various levels
    of pollutants (PCBs, 0.46-7.28 mg/kg dry weight; Cd; Hg) and organic
    matter (5.5-22.3%). During a 100-day exposure period, only small
    increases in PCB concentrations were detected in hard clam
     (Mercenaria mercenaria) and grass shrimp  (Palaemonetes pugio).
    Higher concentrations of PCBs were accumulated by  Nereis virens.
    Uptake was found to be more dependent on the organic content of the
    sediment than on the exposure concentration. Concentration factors
    ranged from 1.59 in a low organic sediment to 0.15 in a high organic
    sediment. The authors also calculated the maximum water exposure
    concentration eluted from each of the sediments. On the basis of a
    concentration factor of 800, calculated by Fowler et al. (1978) for
    the uptake from water of  Nereis sp., body residues of between 0.007
    and 0.034 mg PCBs/kg (wet weight) would have been expected if
    accumulation were dependent purely on direct partitioning from water.
    However, whole-body residues of PCBs were found to be 0.4-0.63 mg/kg,
    suggesting that pathways other than direct uptake from water (e.g.,
    ingestion and sorption) contributed significantly to the accumulation
    of PCBs by the polychaete.

    Freshwater prawns  (Macrobrachium rosenbergii) and clams  (Corbicula
     fluminea) were exposed to contaminated sediments for 48-50 days
    (Tatem, 1986). Prawns were exposed to sediment containing
    approximately 62 mg PCBs/kg (dry weight) and to the same sediment
    diluted with sand to 50 and 10% of the original. Clams were exposed to
    100, 50, or 10% of another sediment containing approximately 2 mg
    PCBs/kg at 100%. The amount of PCBs accumulated was related to the
    exposure concentration, with the highest concentration factors at the
    lowest exposure (10%) level. Bioaccumulation factors for prawns ranged
    from 0.1 to 0.9 for Aroclor 1242 and from 0.2 to 2.4 for Aroclor 1254,
    relative to sediment concentrations. Exposed clams accumulated PCBs
    (Aroclors 1242 and 1254) at concentration factors of 0.54-12.52,
    relative to sediment. When tissues were analysed for Aroclor 1242 and
    1254, maximum concentrations in prawns were attained at 7 and 40 days
    for the 2 Aroclors, respectively. Exposure of prawns at 100 and 50%
    dilution of sediment killed all the animals after 62 and 70 days,
    respectively. Clams survived exposure.

    Clark et al. (1986) investigated the accumulation of sediment-bound
    PCBs by fiddler crabs  (Uca pugilator) and  (Uca minax). Mud and
    mud/sand sediments were used; both were naturally contaminated with
    PCBs and no further PCBs were added. Both species were exposed to a
    mud sediment containing 1.04 mg PCBs/kg and to a mud/sand sediment
    containing 0.37 mg/kg (dry weight). Concentration factors, after a
    28-day exposure, were 0.19 and 0.79, for  U. minax, and 0.2 and 0.59,
    for  U. pugilator, for the 2 sediments, respectively. In a second
    study, using mud with 0.97 mg PCBs/kg and mud/sand with 0.55 mg
    PCBs/kg,  U. pugilator showed concentration factors of 0.58 and 0.71,
    respectively, after 28 days. The authors did not find any detectable
    PCBs in the overlying water, suggesting that the PCBs are tightly
    bound to the sediment and leach out only very slowly. Following
    transfer to uncontaminated sediment on day 42, no PCB residues were
    detected in  U. pugilator on day 56, or in  U. minax on day 63.

    Lynch & Johnson (1982) exposed the amphipod  (Gammarus pseudolimnaeus)
    to 2,4,5,2',4',5'-hexachlorobiphenyl added to sediment in flow-through
    bioassays. Water overflowing from the tank containing the contaminated
    sediment was directed into a second tank where further amphipods were
    exposed without sediment. The hexachlorobiphenyl was labelled with
    14C and added to the sediment at 1 mg/kg; the system was allowed to
    equilibrate for 7-15 days prior to addition of amphipods, which were
    sampled from the tanks after 24, 48, 96, and 192 h. In the initial
    studies, the specific activity of the labelled hexachlorobiphenyl was
    insufficient to detect the hexachlorobiphenyl concentrations in water.
    However, it was clear that amphipods in the tank with the sediment
    accumulated more hexachlorobiphenyl than animals exposed only to the
    water overflow (8.8-10.5 times more PCBs). Removal of organic matter
    from the sediment, by combustion, before addition of the PCB,
    increased uptake of the hexachlorobiphenyl by increasing the
    availability of the material to the  Gammarus. In later studies,
    specific activity was increased and water concentrations could be
    measured. These were very low, ranging between 11 and 35 ng/litre in
    the upper tank and 9 and 25 ng/litre in the lower tank. The lower end
    of this range was found later in the exposure period suggesting that
    less hexachlorobiphenyl was released over time. There was little
    difference in concentration between water taken from the surface and
    that sampled close to the sediment suggesting rapid mixing of the
    overlying water. In this later series of studies, the authors
    demonstrated that both the organic matter content of the sediment and
    the presence of smaller particle sizes (silt and clay) reduced uptake
    of hexachlorobiphenyl by the amphipods. Organic matter was the more
    important factor. Adding maple leaves, to give about 70% organic
    content in the sediment, reduced hexachlorobiphenyl uptake to between
    10 and 20% of that in sediment without organic matter. Very high

    bioconcentration factors were calculated relative to the very low
    water concentrations of hexachlorobiphenyl (ranging between 27 000 and
    1 000 000 in the upper tank and 2000 and 460 000 in the lower tank,
    increasing with exposure period). These factors would be very low
    relative to sediment concentrations of the PCB. However, it is clear
    that the amphipod can accumulate hexachlorobiphenyl, leaching in very
    small amounts from contaminated sediment.

    Cores of lake sediment complete with overlying water were taken by
    Larsson (1984) and transported back to the laboratory, still in the
    sampling tube. PCBs were introduced at different dose levels by
    injection through silicon septa in the walls of the tubes and spread
    evenly 10 mm below the surface. The cores were allowed to stabilize in
    the dark for 1 week at which time 80-100 chironomid larvae were
    introduced. After 8 weeks, the systems were moved and kept at 20C in
    the light. After 2 days, the chironomid larvae began to pupate and
    emerge. The study was terminated after 10 weeks. PCBs were measured in
    sediment, larvae, adults, and exuviae (discarded skins after
    emergence). Ranges of PCBs in sediment were between 0.5 and 14 mg/kg
    giving rise to residues in larvae, exuviae, and adults directly
    related to sediment concentrations. There was "biomagnification"
    between larvae and adult. There was loss of body weight between the
    final larval stage and the adult, but little loss of PCBs (only 17%
    was retained in the exuviae). The author stated that the low variation
    in uptake between animals is an indication of passive physicochemical
    factors being involved in the handling of PCBs by chironomids. Active
    uptake via ingestion would be expected to lead to more variation in
    results. Meier & Rediske (1984) also monitored the uptake of PCBs from
    contaminated sediment into chironomid larvae  (Glyptotendipes
     barbipes). Concentration factors for Aroclor 1242 from sediment
    ranged between 20 and 130 for exposures of between 0.01 and 1.0 mg/kg,
    considerably lower than concentration factors relative to water
    (10 000 for these organisms) (Sanders & Chandler, 1972). Addition of
    oil, commonly found in polluted areas where PCBs spills are likely,
    reduced the uptake of PCBs from the sediment.

    (c) Uptake from food

    A detritus diet containing 17 g Aroclor 1242/kg (wet weight) was fed
    to male fiddler crabs  (Uca pugnax) for 34 days (Marinucci & Bartha,
    1982). The  Spartina detritus was placed in the culture system at the
    start of the study and, because of rapid depletion, was renewed after
    19 days of exposure. Since PCBs leached continually from the food
    source into the water, a second study was carried out to examine the
    uptake of PCBs from water alone. Contaminated detritus was mixed
    thoroughly with water and allowed to equilibrate for 24 h. Water
    levels were found to be 14-15 g/litre. Aroclor 1242 was accumulated
    at a more rapid rate from PCB-laden detritus than from water alone.

    The linear accumulation rate from litter was calculated to be 1 g
    PCBs/day per animal whereas, from water alone, the uptake was 0.1 g
    PCBs/day per animal. Aroclor 1242 was highly concentrated in the
    hepatopancreatic tissue. It was found that the PCB residue in the
    crabs was inversely related to their weight. Comparison of the
    concentrations of PCBs in animals of the same weight shows that, at
    the end of the 34 day exposure, those exposed to water alone had taken
    up approximately half of the PCBs of those exposed to detritus. The
    authors concluded that the crabs in the study accumulated a similar
    amount of PCBs from both the food and the water.

    Pinkney et al. (1985) exposed the amphipod  Gammarus tigrinus to
    Aroclor 1254 (14C-labelled) in fungus  (Fusarium oxysporum) as a
    food item. The fungus contained 195.8 mg Aroclor/kg dry weight.
    Accumulation of PCBs was rapid, reaching a constant level in the
    amphipods of 23 mg/kg after 9-24 h. Similar exposure of the amphipods,
    but with exclusion from direct contact with the fungus by Teflon mesh
    (to monitor uptake of PCBs leached into the water), resulted in
    residues of between 0.16 and 3.3 mg/kg (from concentrations in the
    water at 0.03 g/litre), representing between 0.6 and 13.9% of uptake
    from water and food combined. The PCB residues in the amphipods were
    also monitored over 144 h on uncontaminated food to measure the
    elimination rate. The water was changed every 24 h. Within this
    period, 57% of the accumulated PCBs was eliminated.

    (d) Comparison of different routes of uptake

    In a study by Wyman & O'Connors (1980), the uptake by the marine
    copepods  Acartia tonsa and  Acartia clausi of 14C-labelled Aroclor
    1254 from water, inorganic sediment, and food, was monitored over a
    period of 48 h.  Acartia were exposed to water concentrations of
    10 g PCBs/litre. An asymptotic uptake curve was observed; equilibrium
    was reached after 36 h, corresponding to whole-body residues of 248 mg
    PCBs/kg (dry weight) for  A. tonsa and 223 mg/kg for  A. clausi.
    During exposure, water concentrations fell rapidly to 5 or 6 g/litre.
    A similar pattern of uptake was found after exposure to sediment
    contaminated with 20 mg PCBs/kg with maximum levels of PCBs in
     A. tonsa of 22 mg/kg after 30 h. As in the water exposure, levels of
    PCBs in sediment fell rapidly from 20 mg/kg to 14 mg/kg and then
    slowly to 7 mg/kg at the end of the study. Water levels were initially
    0.62 g/litre and fell to 0.15 g/litre. Uptake of PCBs by  A. tonsa
    from phytoplankton contaminated with 80 mg PCBs/kg (wet weight) was
    very rapid and reached a maximum after 5 h at 61 mg/kg, but
    subsequently declined after exhaustion of the food supply. PCB
    concentrations in water were similar to those found when copepods were
    exposed to contaminated sediment, copepods exposed to these water
    concentrations alone accumulated significantly less PCBs than those
    fed PCB-dosed phytoplankton.

    McManus et al. (1983) exposed the marine copepod  Acartia tonsa to
    14C-Aroclor 1254 either in the food, as phytoplankton containing
    approximately 1.3 mg PCBs, or in water at 1.5 g/litre, for a period
    of 30 h. For copepods exposed to contaminated phytoplankton, PCB
    levels ranged from 117 to 163 mg/kg dry weight. For copepods exposed
    to contaminated water alone, levels ranged from 82 to 104 mg/kg. When
    transferred to clean water, the authors found that copepods lost PCBs
    at a significantly faster rate if they were fed during depuration;
    after 36 h, PCB concentrations in copepods fed during deputation were
    10 mg/copepod whereas those starved contained 30 mg/copepod. No
    significant difference in depuration rate was found between those
    exposed via food and those exposed via water. In a second study,
    elimination in males and females was compared. Although both sexes
    contained similar residues at the start of depuration (117 mg/kg and
    95 mg/kg, respectively), after 36 h, females contained significantly
    lower levels of PCBs than males. During depuration, faecal pellets and
    eggs were analysed; similar levels of PCBs were found in both male and
    female faecal pellets during this period, but levels of PCBs more than
    four times that in the females were found in eggs (407.5 mg/kg dry
    weight after 4 h), indicating that egg production is an important
    route for PCB elimination.  Fish

    Fish of all life stages have been shown to take up PCBs readily from
    water; bioconcentration factors are high. Time taken to reach
    equilibrium is variable, but often long, in excess of 100 days. PCBs
    with greater chlorination are more readily taken up and retained. PCB
    body burden tends to increase with age and levels are higher in fish
    with a greater lipid content. The accumulated PCBs are concentrated in
    lipid-rich tissues. PCBs of lower chlorination are eliminated more
    rapidly. Loss of PCBs is evident when exposure ends; an initial rapid
    loss is followed by a slower rate of loss. Half-life estimates,
    therefore, vary greatly, from a few weeks to several years.
    Reproduction, with the production of a large mass of eggs or sperm,
    allows loss of substantial amounts of the PCB residue. Depending on
    the species, habitat, and behaviour, PCBs can be taken up from water,
    sediment, or food to different degrees.

    (a) Uptake from water

    Califano et al. (1980) maintained larval striped bass  (Morone
     saxatilis) in Hudson river water (filtered and unfiltered)
    contaminated with 14C-Aroclor 1254 at 1.36 g/litre for a period of
    48 h. Whole-body residues for filtered and unfiltered water were not
    significantly different at 5 mg/kg and 5.9 mg/kg, respectively. Uptake
    between 34 and 48 h was very slow, suggesting a steady state had
    already been reached. Exposure of fish for a further 72 h in
    unfiltered water, supported this theory. Elimination was slow, only
    18% being lost in 48 h following a 24-h exposure.

    The PCB uptake pattern in lake trout  (Salvelinus namaycush) sac fry
    was studied by Mac & Seelye (1981) by exposing them to a nominal
    concentration of 50 ng Aroclor 1254/litre for 48 days. Patterns of
    accumulation were similar, regardless of how the data were expressed
    (wet weight, dry weight, or body burden). PCBs levels increased
    slowly, reaching a peak after 32 days (just before completion of yolk
    absorption), and then decreased by day 48.

    Hansen et al. (1975) exposed different life-stages of sheepshead
    minnow  (Cyprinodon variegatus) to Aroclor 1016 (Table 10). After a
    4-week exposure to nominal concentrations of 1, 3.2, or 10 g/litre,
    adult fish laid eggs containing on average 4.2, 17, and 66 mg/kg,
    respectively. DeFoe et al. (1978) exposed fathead minnow  (Pimephales
     promelas) to Aroclor 1248 or 1260 at concentrations of
    0.1-3 g/litre, for 240 days (life cycle). Bioconcentration factors for
    the uptake of PCBs were independent of the PCB concentration in the
    water. Residues in the fish reached an apparent steady state within
    about 100 days of exposure and growth. Females accumulated about twice
    as much PCBs as males, because of their higher body lipid content. The
    variability of residues in females reflected the variability of their
    lipid content. Although mechanisms for uptake were similar for both
    Aroclors, greater body burdens were always achieved with exposure to
    Aroclor 1260. Bioconcentration factors ranged from 60 000 to 160 000
    for males and from 120 000 to 270 000 for females. After transfer to
    clean water, 18% of Aroclor 1248 was lost within 28 days and 15% of
    Aroclor 1260 in 42 days. The authors stated that, because of
    variations between fish, this 10-20% decline in total body burden of
    PCBs was insufficient to indicate definite PCB elimination over this

    De Kock & Lord (1988) exposed an estuarine fish, the Cape stumpnose
     (Rhabdosargus holubi) to a flowing water concentration of 1 g
    Aroclor 1260/litre for 90 days followed by a 90-day period in clean
    water. Equilibrium was reached at 90 days with a concentration factor
    of 24 000. The depuration rate was calculated to be 0.014 days,
    producing a half-life of 50 days.

    Goldfish  (Carassius auratus) were exposed to Clophen A50 at levels
    of 0.01, 0.05, 0.1, or 0.5 mg/litre for 18 days (Hattula & Karlog,
    1973). Rapid uptake was observed with concentration factors of over
    1000 at 18 days, but equilibrium was not achieved within this period.
    Nearly all the fish exposed to 0.5 mg/litre died within 7 days. After
    transfer to clean water, fish that had been exposed to 0.1 mg/litre
    for 13 days and had attained body residues of 70 mg/kg lost half of
    the PCBs within 3 weeks, but still retained levels of approximately
    15 mg/kg, after 70 days.

    Yoshida et al. (1973) exposed carp  (Cyprinus carpio) to 14C-PCBs
    (equivalent to Aroclor 1254) in water or in food. By measuring the
    radioactivity, they found similar tissue patterns of uptake from both
    water and diet. PCBs were localized in the gall bladder, adipose
    tissue, and hepatopancreas and, in particular, the adipose tissue of
    the skull.

    Hansen et al. (1971) exposed spot  (Leiostomus xanthurus) to Aroclor
    1254 at 1 g/litre, for 56 days. Maximum tissue levels of PCBs were
    achieved between days 14 and 28. Highest levels were found in the
    liver (210 mg/kg, after 28 days) followed by the gills, whole fish,
    heart, brain, and muscle. Aroclor 1254 was slowly lost from tissues;
    after 84 days in clean water, levels of PCBs had dropped by 73%.

    In a study by Braun & Meyhofer (1977), rainbow trout  (Salmo
     gairdneri) fingerlings were exposed to water concentrations of 2 or
    20 g Clophen C/litre, for 8 weeks. Tissue PCB concentrations for
    gills, muscle, and liver were found to be 0.62, 0.82, and 3.47 mg/kg,
    respectively, for the lower dose and 12.3, 7.6, and 10.6 mg/kg, for
    the higher dose. When fish were held in clean water for 10 weeks,
    following exposure to 2 g/litre for 8 weeks, residues decreased by
    half in the liver and had disappeared completely from the gills, but
    there was no change in the PCB levels in muscle.

    Rainbow trout  (Salmo gairdneri) were exposed by Guiney et al. (1977)
    to 14C-labelled 2,5,2',5'-tetrachlorobiphenyl at 0.5 mg/litre for
    36 h. The tissue distribution of 14C was measured at regular
    intervals after transfer to clean water. Carcase, muscle, skin, lower
    gastrointestinal tract, and fat contained most of the radioactivity
    (88%). During the first 14 days after exposure, radioactivity
    increased in adipose tissue, carcase, and eyes. Elimination from most
    tissues appeared to be biphasic with a 30% loss within 2 weeks
    followed by a loss of only 6% in the following 126 days. Losses from
    the bile and blood were very rapid and nearly complete within 14 days.
    Based on the initial rate of loss, the authors calculated a half-life
    of 1.55 days, however, the second phase of eliminated PCBs suggested a
    half-life at 2.66 years. In a similar study, Guiney et al. (1979)
    calculated half-lives of 1.76 and 1.43 years for female and male
    rainbow trout, respectively, based on fish sampled 2-34 weeks after
    exposure. For both sexes the half-life of elimination was recalculated
    to 0.52 and 0.54 years between weeks 38 and 52 after exposure (the
    spawning season). The increased elimination appeared to be because of
    loss via eggs and sperm. Vodicnik & Peterson (1985) found a similar
    result after dosing yellow perch  (Perca flavescens); an elimination
    half-life of 22 weeks was calculated. This was later recalculated to
    be <0.7 weeks during spawning, returning to 16.3 weeks after the
    completion of spawning.

    (b) Uptake from sediment

    The uptake of Aroclor 1254 from suspended solids by juvenile Atlantic
    salmon  (Salmo salar) was studied by Zitko (1974). Aroclor 1254 was
    mixed with suspended solids (simulated by SilicAR CC7) in hexane at
    5 mg/ml. Fish were exposed to contaminated solids at 1 g/litre for up
    to 144 days. Over this exposure period, the salmon accumulated 134 mg
    Aroclor 1254/kg.

    Stein et al. (1984) exposed English sole  (Parophrys vetulus) to a
    sediment concentration of 1 mg 14C-Aroclor 1254/kg (dry weight).
    Seawater was allowed to flow over the sediment for 6 days before the
    fish were added. A steady state of PCBs accumulated in the tissues of
    the fish was achieved after 10 days of exposure. Highest residue
    concentrations were found in the bile and the liver. Concentration
    factors were 10 for the bile and 4 for the liver, with other tissues
    individually concentrating PCBs by factors of 3 or less. Simultaneous
    exposure of sole to PCBs and 3H-benzo[ a]pyrene (3 mg/kg, dry
    weight) reduced the amount of PCBs accumulated. Stein et al. (1987)
    collected urban sediment containing aromatic hydrocarbons and PCBs at
    32 mg/kg and 2.2 mg/kg dry weight, respectively. English sole
    accumulated hepatic concentrations of 1.4 mg PCBs/kg (wet weight) over
    a period of 108 days exposure to the urban sediment. This was 8 times
    the PCBs accumulated by sole exposed to the control sediment, which
    did not contain any detectable PCBs. In another study, the same
    authors added a 14C-labelled PCBs tracer to the urban sediment. The
    concentration of PCB-derived radioactivity in the liver reached a
    steady state after 14 days of exposure; the steady state concentration
    in the carcase was found to be significantly lower.

    (c) Uptake from food

    Lieb et al. (1974) fed rainbow trout  Salmo gairdneri on a diet
    containing 15 mg Aroclor 1254/kg for 16 or 32 weeks. PCB levels in the
    lipid fraction increased rapidly for the first 8 weeks, reaching
    equilibrium at about 95 mg/kg. The absolute quantity of PCBs continued
    to increase as the fish grew. The trout had retained 68% of the total
    PCBs ingested at equilibrium. No elimination was found after transfer
    to uncontaminated food at 16 weeks (for a period of 16 weeks), or
    after starving the fish for 8 weeks following exposure for 32 weeks.
    Reductions in PCB levels were found, but these were cancelled out by
    concomitant reductions in lipid content.

    Coho salmon  (Oncorhynchus kisutch) parr were fed 10 mg chloro-
    biphenyls/kg (containing equal parts by weight of 3,4,3',4'-tetra-
    chlorobiphenyl, 2,4,5,2',4',5'-hexachlorobiphenyl, and
    2,4,6,2',4',6'-hexachlorobiphenyl) for up to 165 days (Gruger et al.,
    1975). Most of the PCBs were accumulated in the adipose tissue of the
    salmon (51.1 mg/kg total chlorobiphenyls after 165 days). Tissue

    levels of tetrachlorobiphenyl were about half those of either of the
    two hexachlorobiphenyls throughout the exposure period. When fish were
    starved for 48 days, the data indicate mobilization or transformation,
    with, for example, chlorobiphenyls in the spleens lowered by half and
    in adipose tissue increased 5-fold. Most tissues showed an increase in
    PCB levels, especially blood levels. In contrast, when a second group
    of salmon were fed on a clean diet, chlorobiphenyls were released from
    adipose tissue and levels increased in some other tissues, such as the
    lateral line dark muscle tissue. The ratio of the different
    chlorobiphenyls remained unchanged during both of these post-exposure
    treatments. Gruger et al. (1976) fed juvenile coho salmon diets
    containing a mixture of 2,5,2',5'-tetrachlorobiphenyl, 2,4,5,2',5'-
    pentachlorobiphenyl, and 2,4,5,2',4',5'-hexachlorobiphenyl
    at 1, 2, and 12 mg/kg, for up to 72 days. A steady state appeared to
    have been reached between 17 and 35 days at the lowest dose (a whole
    body concentration of approximately 0.45 g/kg (wet weight)); steady
    state was not achieved at the other 2 dose levels. All 3
    chlorobiphenyls were accumulated to similar levels. Comparing these
    data with the study by Gruger et al. (1975), suggests that the
    position of the chlorine substitution is an important factor.

    Hansen et al. (1976a) fed channel catfish  (Ictalurus punctatus) on a
    diet contaminated with 20 mg Aroclor 1242/kg. The total burden of PCBs
    (g PCB/fish) increased exponentially with exposure time. When fish
    were placed on a clean diet (from day 84 for 56 days) a slight net
    decrease in body burden was observed, but levels remained constant
    when fish were placed on a clean diet for 56 days after 140 days
    exposure. On return to a PCB-contaminated diet, accumulation rates
    returned to those previously observed. The authors noted that, during
    PCB-free periods, there was a shift in residues from edible carcase to

    Mayer et al. (1977) fed fingerling coho salmon with Aroclor 1254 at
    concentrations ranging between 1.45 and 14 500 g/kg body weight.
    Equilibrium was reached after 112 days at concentrations of 1.45,
    14.5, and 145 g/kg, with whole body residues of 0.47, 0.5, and
    3.8 mg/kg, respectively. A steady state was reached at the 2 highest
    dose levels of 1450 and 14 500 g/kg after 200 days, with
    corresponding residues of 57 and 659 mg/kg. In another study, channel
    catfish  (Ictalurus punctatus) were exposed to Aroclors 1232, 1248,
    1254, and 1260 in the diet at concentrations of 48 or 480 g/kg body
    weight, for 193 days. Equilibrium was only achieved at the lowest
    exposure dose of Aroclor 1232, within 150 days, with a whole-body
    burden of 4.5 mg/kg. Similar whole-body residues were achieved at the
    lowest dose of the other Aroclors, but no steady state was reached. At
    the higher dose, accumulation increased in the order Aroclor 1232 =
    1248 < 1254 < 1260, with residues ranging from 13 to 32 mg/kg after
    193 days.

    When Zitko (1974) fed juvenile Atlantic salmon  (Salmo salar) diets
    containing 10 or 100 mg Aroclor 1254/kg, accumulation reached
    equilibrium within 30 days at the lower dose, with a whole-body
    residue of approximately 3.8 mg/kg. Equilibrium was not reached within
    200 days at 100 mg PCBs/kg. A whole-body residue of 30 mg/kg was
    recorded at 181 days.

    Zinck & Addison (1974) administered a mixture of 2-, 3-, and
    4-chlorobiphenyl to thorny skate  (Raja radiata) and winter skate
     (Raja ocellata) by intravenous injection. All three congeners were
    cleared rapidly from blood plasma, 3-chlorobiphenyl consistently being
    cleared more rapidly than the other two. Less than 6% of
    3-chlorobiphenyl remained in the plasma after 15 min compared with 30%
    for the other chlorobiphenyls. All three accumulated in the other
    tissues of  R. radiata, principally in the liver and muscle. Tissue
    levels of 3-chlorobiphenyl were consistently less than the others
    during the 53-h sampling period.

    In a study by Guiney & Peterson (1980), both yellow perch (a non-fatty
    fish) and rainbow trout (a fatty fish) were dosed with 0.8 g of
    14C-labelled 2,5,2',5'-tetrachlorobiphenyl, either orally or by
    intraperitoneal injection. Whole-body elimination was found to be
    similar for both species and routes. A 20-30% elimination was observed
    after 3-4 days with virtually no more PCBs being eliminated during the
    rest of the 32-day sampling period. Tissue distribution varied between
    the 2 species; uptake in the perch was mainly concentrated in the
    viscera and carcase, whereas, in the trout, skeletal muscle and
    carcase were the major sites of uptake.

    Niimi & Oliver (1983) calculated the biological half-life of 31
    dichloro- to decachlorobiphenyl congeners, 105 days after a single
    oral dose of 46-261 mg/kg was administered to rainbow trout  (Salmo
     gairdneri). Whole-body half-lives increased from 5 days to >1000
    days as the number of chlorines on the biphenyl increased. From
    structure-activity analysis of half-lives in whole fish, the authors
    concluded that elimination was enhanced for congeners with a lower
    chlorine content and no chlorine substitutions in the  ortho
    positions, and for those with 2 unsubstituted carbons adjacent on the
    biphenyl.  Birds

    PCBs are taken up from contaminated food or water and concentrated in
    the fatty tissues of birds. PCBs of higher chlorination levels are
    accumulated to a greater extent. Egg-laying females can lose
    substantial amounts of PCBs from body tissues by transferring the PCBs
    to the eggs. Redistribution of residues occurs on starvation (of

    significance during the migration of birds in the wild). Expressed as
    a whole-body concentration, PCB residues fall during starvation.
    However, expressed as a concentration in fat, residues rise. Most
    critically, PCB residues in the brain increase during starvation and
    this may kill the birds without further intake of PCBs.

    Brunstrm et al. (1982a) injected the yolk of developing hens' eggs,
    on day 4 of incubation, with 14C-labelled 2,4,2',5'-tetra-
    chlorobiphenyl at a concentration of 5 mg/kg. One hour after
    injection, radioactivity was found in the sub-blastodermic fluid, the
    highest concentrations being in amniotic membranes. None was present
    in the yolk, albumen, or embryonic tissues. Uptake was uniform
    throughout the embryo, after one day, and, as tissues developed,
    became concentrated in certain of them, such as the liver, kidney, and
    fluid brain vesicles, by day 7. 14C was found uniformly in the yolk
    after 11-14 days and was highly concentrated in the first bile
    produced on day 11. The labelled PCBs accumulated in fatty tissue as
    it developed from day 14 onwards. In the hatched chick, large amounts
    of radioactivity were found to be concentrated in the gall bladder,
    intestine, cloaca, and the coiling of the gizzard. When either
    3,4,3',4'-tetrachlorobiphenyl or 2,4,2',5'-tetrachlorobiphenyl was
    injected into the air sac of hens' eggs on day 14 of incubation at
    0.4 mg/kg, no difference in distribution pattern was observed 1-5 days
    later (Brunstrm & Darnerud, 1983). The highest amounts of
    radioactivity were found in the fatty tissue, liver, kidneys, and the
    gall bladder, 14C was also found in the bone marrow, the adrenals,
    and the gonads, but to a lesser extent. The yolk contained less
    radioactivity than the yolk analysed in the previous study by
    Brunstrm et al. (1982a), because the PCBs were administered via the
    air sac.

    White leghorn hens were exposed to 50 mg Aroclor 1254/litre in their
    water for 6 weeks (Tumasonis et al., 1973). PCB residues in the yolks
    of eggs laid increased during the exposure period to a peak, after 6
    weeks, of approximately 205 mg/kg. When hens were given clean water,
    the yolk levels of PCBs quickly dropped within 5 weeks to
    approximately 100 mg/kg, and then more slowly until, after 20 weeks
    without Aroclor 1254 in their water, the hens laid eggs containing
    0.7 mg/kg.

    During a 4-week exposure to Aroclor 1242, 1254, or 1260, in the feed
    of one-day-old chicks, Harris & Rose (1972) found that PCBs
    accumulated in the fat and that this accumulation increased with
    increasing exposure concentrations of 100, 200, and 400 mg/kg. At the
    2 highest dose levels, the hens accumulated more of Aroclor 1260 than
    of the other 2 Aroclors (i.e., 482, 1427, and 2151 mg Aroclor 1260/kg
    at the 3 exposure concentrations, respectively). At the highest dose,
    there was high mortality during exposure to Aroclor 1242 and 1254 and
    this might have affected the residues found.

    Greichus et al. (1975) fed white pelicans  (Pelecanus erythrorhynchos)
    on a fish diet containing 100 mg Aroclor 1254/day, for 10 weeks. PCB
    residues were measured in the carcase, liver, feathers, and brain;
    mean residues found were 2130, 290, 120, and 110 mg/kg wet weight,

    In a study by Dahlgren et al. (1972), 11-week-old pheasant  (Phasianus
     colchicus) were dosed with one capsule per day containing 210 mg of
    Aroclor 1254. Birds that died between days 1 and 5 contained, on
    average, PCB residues of 520 mg/kg in the brain, 2500 mg/kg in the
    liver, and 140 mg/kg in muscle. Birds that were sacrificed over the
    same period had mean brain, liver, and muscle PCB levels of 370, 1900,
    and 83 mg/kg, respectively. All birds dosed with only 10 mg of Aroclor
    1254 per day died within 180 days and contained average brain and
    liver residues of 360 and 1200 mg/kg, respectively.

    Sdergren & Ulfstrand (1972) fed robins  (Erithacus rubecula)
    mealworms containing 1 g of Clophen A50/day for 15 days. Brain,
    breast muscle, and carcase were analysed and contained mean PCB
    residues of 0.35, 0.55, and 4.5 mg/kg fresh weight, respectively. A
    second group of robins was starved following dosing and all died
    within 48 h. PCB levels were higher in the brain and breast muscle at
    1.1 and 1.3 mg/kg, respectively, but carcase PCB levels were lower on
    a fresh weight basis at 2.6 mg/kg. When the carcase lost some of its
    fat content during starvation, PCB levels in terms of fresh weight
    decreased. Consequently, because of the low remaining fat content,
    residue levels in terms of fat weight increased. Another group of
    birds were fed both PCBs and DDT (10.5 g/day) for 15 days and then
    starved. PCB levels in all 3 tissues analysed were higher than those
    in birds administered PCBs alone followed by starvation; residues
    were: brain, 9.3 mg/kg fresh weight, breast muscle, 8.8 mg/kg, and
    carcase, 4.5 mg/kg.

    Cormorants  (Phalacrocorax carbosinensis) were kept on a fish diet
    contaminated with PCBs for one month, followed by gelatin capsules of
    PCBs administered daily for the remainder of the exposure (Koeman et
    al., 1973). After 14 weeks, the dose rate of Clophen A60 was increased
    periodically during the exposure period from 200 to 500 mg/kg. The
    birds died between days 55 and 124, and overall residues of PCBs
    increased in the tissues, the longer the birds survived. Total-body
    residues ranged from 850 to 2750 mg PCBs/kg (wet weight) at death.
    Brain and liver residues ranged from 76 to 180 mg/kg and from 210 to
    290 mg/kg, respectively. The fat of 2 birds was analysed for PCBs and
    was found to contain 10 300 and 20 500 mg/kg.

    Harris & Osborn (1981) dosed wild puffins  (Fratercula arctica) by
    implantation with 30-35 mg of Aroclor 1254. PCBs were quickly taken up
    in fat, with concentrations rising to 10-14 times that in control
    birds (highest fat residue 654 mg/kg wet weight), and remaining at
    this level for up to 10 months. Levels slowly declined, but were still
    twice those of controls after 34 months. PCB concentrations in the
    liver and muscle tissue were highest shortly after dosing (48.4 and
    25.2 mg/kg, respectively) and declined until, after 16 months, no PCBs
    were detectable. Levels of PCBs in the kidneys and brain were variable
    with no consistent trends.

    Common grackles  (Quiscalus quiscula), starlings  (Sturnus vulgaris),
    red-winged blackbirds  (Agelaius phoeniceus), and brown-headed
    cowbirds  (Molothrus ater), were fed diets containing 1500 mg Aroclor
    1254/kg over an 8-day period (Stickel et al., 1984). PCB residues in
    the brains of birds that died were found to be higher than those in
    birds that were sacrificed over a similar period. PCB residues ranged
    from 349 to 763 mg/kg in birds that died and from 54 to 301 mg/kg in
    birds sacrificed. Liver and whole-body residues tended to be higher in
    birds that died, but they overlapped to a large extent. PCB residues
    in whole bodies on a lipid basis showed the most clear-cut difference,
    ranging from 22 600 to 98 600 mg/kg for birds that died and from 6690
    to 22 500 mg/kg for those sacrificed. PCB residues in grackles
    declined slowly, when the birds were placed on a clean diet. From a
    whole-body level of 1300 mg/kg, residues declined to 169 mg/kg, 224
    days later. The rate of decline was irregular, but a half-life was
    estimated at 89 days over this period of loss.  Mammals

    Olsson et al. (1979) fed mink  (Mustela vison) on a diet containing
    11 mg PCBs/kg for 66 days. Mink accumulated 310 mg PCBs/kg in
    extractable fat over the exposure period. Control mink were found to
    contain 14 mg PCBs/kg, and, when the control feed was analysed, it was
    found to contain 0.05 mg PCBs/kg. The authors also found a significant
    increase in cadmium uptake in the kidneys of PCB-treated animals
    compared with controls. In another study on mink  (Mustela vison),
    Hornshaw et al. (1983) administered various PCB-contaminated fish
    diets containing between 0.21 and 1.5 mg PCBs/kg. Adipose tissue
    samples were taken after 6-8 weeks and after 18 weeks exposure (Table
    11). The amount of PCBs accumulated was directly related to the amount
    of PCBs in the diet; mean PCB residues ranging from 4 to 24.8 mg/kg
    after 6-8 weeks and from 8.1 to 42.8 mg/kg after 18 weeks. When
    expressed as individual congeners, it can be seen that the mink showed
    the highest accumulation of the PCBs with the chromatographic peak
    corresponding to 2,4,5,2',4',5'-hexachlorobiphenyl. To determine the
    rate of PCB elimination, male mink that had been on a fish diet

    containing 1.5 mg PCBs/kg for 10 weeks were transferred to a control
    diet. Over this period, adipose tissue residues of 32 mg PCBs/kg had
    accumulated. Over the 16-week elimination period, 60.3% of the total
    PCB burden of the adipose tissue was eliminated. This consisted of a
    loss of 87.2% of 2,5,2',5'-tetrachlorobiphenyl, 88.9% of
    2,3,6,2',5'-pentachlorobiphenyl, and 55.4% of the hexachlorobiphenyl.
    The half-life for total PCBs in mink adipose tissue was calculated to
    be 98 days.

    Wren et al. (1987a,b) fed mink on a commercial mink food supplemented
    with 1 mg Aroclor 1254/kg for a period of 6 months. Male mink had
    liver residues of 1.98 mg PCBs/kg after 118 days and 2.8 mg/kg after
    183 days exposure. The liver of a female, analysed on day 161
    contained a residue of 3.1 mg PCBs/kg. Liver PCB levels in 5-week-old
    kits were similar to those in adult mink fed the experimental diet for
    several months. Bleavins et al. (1981) measured the relative
    importance of placental transfer and milk in the transfer of PCB
    residues from mother mink to offspring. Newborn kits contained less
    than 0.1% of a dose of PCBs injected into the mother mink. At 2 weeks
    of age, the kits contained 1.2% of the dose given to the mother,
    suggesting that lactation is a major route of exposing the young to
    PCBs and a major route for the loss of PCBs from the mother. Placental
    transfer of PCBs was greater in the ferret than in the mink (Bleavins
    et al., 1984). The ratio of placental to mammary transfer was 1:15 for
    offspring whose mothers were dosed during the first trimester of
    pregnancy and 1:7 for mothers exposed during the last trimester.

    Big brown bats  (Eptesicus fuscus) were fed on mealworm diets
    containing 9.4 mg Aroclor 1254/kg for up to 37 days (Clark & Prouty,
    1977). In bats sacrificed on day 37, residues ranged from 29 to 121 mg
    PCBs/kg (wet weight) for the carcase and from not detectable to
    4.2 mg/kg in the brain. Bats that were starved following exposure
    showed a significant correlation between increasing brain PCB
    concentrations and carcase lipid concentrations. The authors stated
    that PCBs increased in brain tissue as carcase fat was metabolized.
    Clark (1978) exposed pregnant big brown bats to a mealworm diet
    containing 6.36 mg Aroclor 1260/kg for approximately 18-28 days, until
    the young were born. Mean carcase levels of PCBs were 20.34 mg/kg in
    parent females and 4.38 mg/kg in litters. Levels of PCBs in both
    adults and young continued to rise throughout the sampling period; the
    longer the gestation time, the higher the PCB level in the sample.

    4.2.5  Appraisal

    Experimental work on mammals has been concentrated on terrestrial
    species. Problems with PCB toxicity are important for marine mammals,
    but these are less convenient for experimental study. Results in this
    section, therefore, have to be related to field observations on marine

    Mink take up more chlorinated components of PCB mixtures and can
    accumulate large residues of PCBs. On cessation of exposure, more
    tetrachloro- and pentachlorobiphenyls were eliminated than
    hexachlorobiphenyl. The half-life for total PCBs was calculated to be
    98 days. PCB residues are transferred from mother to offspring. The
    relative importance of transplacental transfer and transfer in milk
    varies between species. Redistribution of residues takes place on
    starvation, which is of significance for migratory species; brain
    residues, which may be fatal with no further intake of PCBs, increase
    as animals are starved.


    5.1   Levels in the environment

    PCBs were detected in the environment in the late 1960s (Risebrough et
    al., 1968; Jensen et al., 1969) and, within a short time, were
    reported as contaminants in almost every component of the global
    ecosystem including air, water, soil, fish, wildlife, human blood,
    adipose tissues, and milk (Holdrinet et al., 1977; Wassermann et al.,
    1979; Ballschmitter et al., 1981; Buckley, 1982; Safe, 1982b; Bush et
    al., 1985; Kannan et al., 1988; Tanabe, 1988).

    The lipophilic properties of PCBs are the basis of the bioaccumulation
    and biomagnification that has been demonstrated and, thus, numerous
    sources within the environment can lead to human exposure.

    High-resolution, gas-chromatographic analysis has shown that the
    congener composition and relative concentrations of the individual
    components in many PCB extracts from environmental samples differ
    markedly from those in the commercial PCBs (Jensen & Sundstrm, 1974a;
    Wolff et al., 1982a; Safe et al., 1985a; Brown et al., 1987a,b).

    A major problem with data concerning PCB levels in environmental
    samples is that they normally are only available for "total PCBs" and
    that there are much fewer data on actual "PCB patterns". Moreover,
    when comparing results produced from different laboratories or from
    the same laboratory at different times, an additional difficulty may
    arise from differences in the sampling and analytical techniques used.

    It is difficult, if not impossible, to compare data obtained with
    different analytical methods, from different laboratories, and
    countries. Nowadays, the older data seem less reliable, especially in
    the light of the use of improved analytical methods and better
    sampling techniques (WHO/EURO, 1987). A comprehensive review of world
    PCB levels was published by Wassermann et al. (1979).

    5.1.1  Air

    PCB concentrations in air differ markedly from location to location,
    with the lower levels found over the oceans or over non-industrialized
    regions, such as the Canadian Northwest territories. In general,
    levels over industrialized areas or over landfills are the highest.
    Apparently, these levels influence PCB levels in rainwater and there
    is a gradient of values in air from industrial to rural areas. Some
    typical values can be found in Table 12.

    MacLeod (1981) described a method for the analysis of PCBs using
    low-volume, indoor air sampling to estimate the presence of PCBs in
    indoor air in work-places and homes in the USA. Three facilities, an
    industrial research facility, an academic facility, and a shopping
    complex were sampled. The periods of sampling ranged from 2 days up to
    6 months. The average concentrations (calculated as Aroclor 1242 plus
    Aroclor 1254) ranged from 44 up to 240 ng/m3. Outdoor levels of up to
    18 ng/mg3 were found. In the homes, air samples from 14 areas (of
    which 9 were kitchens) were also analysed. The average concentrations
    in the kitchens ranged from 150 up to 500 ng/m3 and, in the other
    rooms, from 39 to 170 ng/m3. In a library, a level of 400 ng/m3 was

    The levels of PCB exposure that may occur in buildings in the USA were
    determined by Oatman & Roy (1986). Air samples and surface wipe
    samples were taken in 5 state-owned, office buildings and 2 elementary
    schools. The average levels of airborne PCBs in buildings with PCB
    transformers were nearly twice the levels in buildings without
    transformers, i.e., 457  223 and 229  106 ng/m3, respectively. The
    mean of the surface wipes taken in buildings without PCB transformers
    was 0.17 and that in the buildings with transformers 0.23 g/100 cm2.
    There was a wide variation between the different buildings and, as
    shown above, the presence of transformers influenced the indoor PCB
    concentrations.  Rain and snow

    In the Netherlands, at Bilthoven, the PCB-concentrations in rainwater
    ranged from 0.01 to 1.5 g/litre (van Zorge, cf. WHO/EURO, 1985). In
    the Federal Republic of Germany, concentrations of 0-4 ng/litre were
    found (DFG, 1988).

        Table 12.  PCB levels in air in several countries

    Country      Location and/or type of sample                                  PCB levels               References
                                                                                 average and/or range

    Canada       Northwestern territories                                        0.002-0.07 ng/m3         Bidleman et al. (1978)

    Germany      Industrial area (Ruhr area)                                     3.3 ng/m3                DFG (1988)
                 Non-contaminated area                                           0.003 ng/m3

    Japan        Within industrial plants:    - PCB vapours                      13-540 g/m3             Tatsukawa & Watanabe
                                              - PCBs on airborne particulates    4-650 g/m3              (1972)

                 North Pacific, South Pacific, Indian,                           0.1-0.3 ng/m3            Tatsukawa & Tanabe
                 Antarctic and South Atlantic Oceans                                                      (1983)

                 North Atlantic Ocean                                            0.5 ng/m3                Tatsukawa & Tanabe

    Sweden       Several locations                                               0.8a-3.9 ng/m3           Ekstedt & Odn (1974)

    USA          Near the North-East Coast                                       5 ng/m3                  Harvey & Steinhauer
                 Over the Atlantic Ocean, 2000 km away                           0.05 ng/m3               (1974)
                 from the industrial complex

                 several locations                                               1-50 ng/m3               Panel on Hazardous
                                                                                                          Substances (1972)
                                                                                                          cf WHO/EURO (1988)

    Yugoslavia   Bela Krajina:                - 300 m from an industrial plant   4-7 g/m3                Jan et al. (1988b)
                                              - air near a waste landfill        45 g/m3
                                              - over the River Kruga             2-5 g/m3

    a  Limit of determination.  Natural gas

    PCBs were first identified in gas pipelines in January 1981, when a
    PCB-containing oil condensate was found in the gas meters of some
    residential customers in Long Island, New York. Voluntary monitoring
    of condensate and natural gas by 33 transmission companies, showed the
    presence of PCBs in 12 companies. PCBs were also found in gas
    pipelines. Condensate is a mixture of heavier hydrocarbons and other
    liquids, such as water, that condenses, because the gas is transmitted
    under pressure. This condensate tends to collect in pools in the
    pipes. In the period 1981-83, 1841 samples of condensate from gas
    pipelines were analysed: 659 (35.8%) of the samples contained
    < 25 mg/kg; 65.8% of the samples contained <1000 mg/kg, and 0.4%,
    > 10 000 mg/kg. The maximum level that was found was 42 394 mg/kg
    (Versar Inc., 1984).

    In the period 1981-83, 138 samples of natural gas in transmission were
    analysed. In 29 samples, PCBs were found with a minimum concentration
    of <0.004 g/m3 and a maximum concentration of 1050 g/m3. Natural
    gas in distribution lines was also analysed in the same period. Out of
    528 samples, 224 did not contain any PCBs. The levels ranged from
    <0.02 to 51 g/m3.

    Indoor concentrations (kitchens, etc.) were measured in 419 samples in
    the period 1981-83. No PCBs could be detected in 49 samples, but, in
    the others, levels ranged from <0.01 to 1.08 g/m3 (Versar Inc.,

    5.1.2  Water

    Surface water may become contaminated with PCBs from atmospheric
    fall-out or from direct emissions from point sources. Because of
    adsorption on suspended particles, PCB concentrations in heavily
    contaminated waters may be several times greater than their
    solubility. Sdergren (1973) reported a seasonal variation, which was
    attributed to aerial fall-out.

    It has been shown that polluted rivers, lakes, and estuaries have
    higher PCB values than non-polluted waters (Table 13). On the basis of
    scanty information on PCBs and reinforced by extensive analogue
    information on DDT, it has been estimated that, for the Great Lakes of
    North America, non-polluted freshwaters might contain less than
    5 ng/litre, moderately polluted rivers and estuaries, 50 ng/litre, and
    highly polluted rivers, 500 ng/litre. These values can be used to
    evaluate those reported by several authors and presented in Table 13.

        Table 13.  PCB levels in water in several countries

    Country       Location and/or type of                                   PCB levels           References
                                                                            sample               average and/or range

    Germany       Several rivers                                            5-103 ng/litre       Lorenz & Neumeier (1983)

    Netherlands   River Rhine (1976/1977)                                   100-500 ng/litre     Wegman & Greve (1980)

    Sweden        Water entering a treatment                                0.5 ng/litre         Ahling & Jensen (1970)

                  Tap water produced at the plant                           0.33 ng/litre        Ahling & Jensen (1970)

                  Several rivers                                            0.1-0.3 ng/litre     Ahnoff & Josefsson (1974)

    USA           Polluted coastal area                                     100-450 ng/litre     Panel on Trace Hazardous
                  Lake Michigan (1970)a                                                          Substances (1972) (cf.
                                                                                                 WHO/EURO, 1988)

                  Distribution system feeding the Fort Edwards reservoirs   < 12-160 ng/litre    Brinkman et al. (1980, 1981)
                  in New York (1978)

                  Hudson River at Fort Edward                               up to 530 ng/litre   Brinkman et al. (1980, 1981)

    a  Followed by a marked decrease in 1971.

    5.1.3  Soil

    Soil may become contaminated with PCBs from atmospheric fall-out or
    from direct emissions from point sources. The presence and behaviour
    of these compounds in the soil depend on substance (congener)-specific
    characteristics and on a number of soil parameters. Sorption and
    condensation processes in the soil also play a role in the removal of
    PCBs. Some values of PCB levels in soil can be found in Table 14.

    Klein (1983) found that PCBs accumulate in the sediments of rivers and
    lakes in the Federal Republic of Germany and that these levels
    indirectly reflect the contamination of water by PCBs. Some values for
    PCBs in sediments can also be found in Table 14.

    An important, though localized, source of PCB contamination of soil,
    can be the use of sewage sludge as a fertilizer in agriculture. PCB
    levels varying from 0.1 to 765 mg/kg (dry weight) have been reported
    in sewage sludge from different countries, the usual range being 0.1
    to 9.0 mg/kg (WHO/EURO, 1987). In the USA, 16 sewage sludge samples
    from cities contained a mean Aroclor 1254 concentration of 5.2 mg/kg
    dry weight (range 0.01-23.1 mg/kg). Other authors reported a range of
    1.5-27.3 g/litre in 36 raw sewage sludges. Some levels that have been
    found for PCBs in sludges are presented in Table 14 (WHO/EURO, 1987).

    Five sediment samples were collected from the Waukegan Harbour of Lake
    Michigan, Illinois, in 1978. Residues of 3,4,3',4'-tetrachlorobiphenyl
    ranged from 0.005 to 27.5 mg/kg and residues of 2,3,4,3',4'-penta-
    chlorobiphenyl, from 0.102-131 mg/kg. The total PCB contents of the
    sediment ranged from 10.6 to 13 360 mg/kg (Huckins et al., 1988).

    5.1.4  Aquatic and terrestrial organisms

    PCBs have been measured in a wide variety of biota from many different
    locations throughout the world. Only a few illustrative examples are
    given here, more comprehensive lists of PCB residues can be found in
    reviews by Risebrough et al. (1968); Peakall (1975); and Eisler
    (1986). Tanabe et al. (1987) reported that the highly toxic, coplanar
    PCBs are as widely spread as general PCB pollution.

    In the biota of a small upstate New York public water supply system,
    which is near the polluted section of the Hudson River and a disposal
    site of PCB-containing waste, PCBs were found in detectable
    concentrations (Table 13). Five samples of algae showed Aroclor 1254
    levels of <25 (nd)-120 g/kg dry weight, macro-invertebrates showed
    levels between <200 and 3800 g/kg and vertebrates, between <25 and
    1100 g/kg dry weight (Brinkman et al., 1980, 1981).

        Table 14.  PCB levels in soils, sediments, and sewage sludge in several countries

    Country          Location and/or type of sample                            PCB levels               References
                                                                               average and/or range

    Germany          Soil without sewage sludge                                0.02-0.08 mg/kga         Markard (1988)
                     Soil with sewage sludge                                   0.05-3.0 mg/kga
                     Sewage sludge                                             ndb-19 mg/kg
                     Sediments of contaminated waters                          0.1-1.0 mg/kga           Klein (1983)
                     Sediments of several rivers                               0.16-0.59 mg/kg          DFG (1988)
                     Agricultural soil                                         0.03 mg/kg               DFG (1988)

    Japan            Agricultural soil                                         < 1 mg/kg                Fukada et al. (1973)
                     Soil near a factory making electrical components          510 mg/kg                Fukada et al. (1973)

    Netherlands      Sediments from several surface waters                     < 0.01-1.2 mg/kga        Greve & Wegman (1983)

    United           Soil from a waste disposal area with chemical treatment   4.5-44.8 g/kg           Eduljee et al. (1986);
    Kingdom          and incineration facilities                                                        Badsha et al. (1986);
    (Scotland)       Grass samples from the same area (foliage)                2.9-64.7 g/kg           Badsha & Eduljee (1986)
                     Soil of rural areas                                       8 g/kga (1-23)
                     Grass of rural areas                                      9 g/kga (7-16)
                     Soil of urban areas                                       52 g/kg (11-141)
                     Soil of industrial locations                              41 g/kg (20-67)

    United           Surface soil                                              2.5 g/kg                Jones (1989)
    Kingdom                                                                    (0.2-12.2)

    Table 14.  (cont'd).

    Country          Location and/or type of sample                            PCB levels               References
                                                                               average and/or range

    USA              Sediments near a point of accidental release of PCBs      1.4-61 mg/kg             Nimmo et al. (1971a)
    Escambia         Sediments 16 km downstream of this point                  0.6 mg/kg
    Escambia         Soil samples from the bank, 6.5 km downstream from the    1.4-1.7 mg/kg
    Bay              point

    a  Dry weight
    b  Not detectable

    Serious environmental contamination has been documented in enclosed
    water bodies close to urban and industrialized areas, such as the
    Great Lakes, the Baltic Sea, and Tokyo Bay. PCB levels in aquatic
    organisms reflect these localized high concentrations.

    Nimmo et al. (1971a) reported that PCB levels in shrimp from Escambia
    Bay, Florida (contaminated by an industrial plant on the Escambia
    River) contained between 0.6 and 120 mg Aroclor 1254/kg in 1969 and
    fiddler crabs, collected in 1970, contained 0.45-1.5 mg/kg.

    When fish, sampled throughout the USA, were analysed by Schmitt et al.
    (1983, 1985), the highest levels of PCBs were found in the
    North-eastern industrialized areas. Delfino (1979) reported
    concentrations ranging from 26 to almost 1000 mg PCBs/kg in fish
    collected from the Sheboygan River, Wisconsin, contaminated by a
    die-casting plant.

    Wiemeyer et al. (1975) analysed osprey eggs in 1968-69 and found
    average levels of 2.6 mg/kg in Maryland compared with an average level
    of 15 mg/kg in eggs from Connecticut. PCB residues in Connecticut eggs
    had not changed significantly compared with those collected in 1964.

    Buckley (1982) analysed aspen, sumac, and golden rod plants growing at
    various distances (< 1200 m) and in different directions from a PCB
    dump in New York State, USA. All the plants were growing beyond a
    natural drainage ravine, which prevented contamination of soil and
    water by PCBs. Downwind of the site, PCB levels in the plants were
    found to be approximately 100 mg/kg dry weight (over 600 times
    background levels in plants). Levels above background concentrations
    were also found in directions from the site less obviously
    contaminated by airborne dust.

    Eggs of terrestrial birds collected in a rural environment in Canada
    contained lower PCB levels than those sampled from urban areas (Frank
    et al., 1975).

    In the Great Lakes, the highest levels of PCBs were found in Lakes
    Michigan and Ontario for fish (Delfino, 1979) and Lake Ontario for
    birds (Weseloh et al., 1979); both lakes receive input from industrial
    and urban sites. Glooschenko et al. (1976) found concentrations of up
    to 8.1 mg/kg in microorganisms from the middle of Lake Huron.

    Weseloh et al. (1983) found that the PCB levels in double-crested
    cormorant eggs, collected from Lake Superior during 1972 (average of
    23.8 mg/kg fresh weight), were higher than those in cormorant eggs
    analysed in other Canadian colonies. Mineau et al. (1984) found that
    the locations of herring gull colonies with the greatest mean levels
    of PCBs, in each of the Great Lakes, corresponded with the locations
    of major sources of the contaminant, as indicated by elevated residues
    in sediment.

    Muir et al. (1988) determined PCB levels in pooled Arctic cod muscle
     (Boreogadus saida) and polar bear fat  (Ursus maritimus), and in
    the blubber and liver of ringed seals  (Phoca hispida) from 3
    locations in the East/Central Canadian Arctic. The mean arithmetic
    concentrations of total-PCBs in the muscle of Arctic cod of 2
    locations were 3 and 5 g/kg wet weight. The mean concentrations shown
    in the tabulation below were found in the blubber and liver of ringed


    Year        Number of     Sex          Arithmetic mean  SD
                  samples                  (g/kg wet weight)

    1972              3       female        639  249
    1975/76           5       female        600  99
    1983             10       male          794  879
                     16       female        308  138
    1984             19       male          568  287
                     14       female        375  172
    1984             19       male            6  4
                     14       female          4  3

    The presence of PCBs in polar bears  (Ursus maritimus) was studied by
    Norstrm et al. (1988) in the Northwest territories of Canada. Liver
    and adipose tissue specimens were obtained by Inuit hunters from 12
    zones over the period 1982-84. A total of 121 samples was obtained.
    The mean concentrations of total PCBs in pooled samples ranged from
    3.24 to 8.25 mg/kg, on a lipid weight basis. The adipose tissue of
    polar bear (10 pooled samples collected in 1982 and 10 samples, in
    1984) contained 4.42 and 4.57 mg/kg wet weight, respectively. From
    these results, biomagnification factors for the food-chain of the
    Arctic cod/ringed seal/polar bears were calculated. For total PCBs,
    these factors ranged from 3.7 to 8.8 for fish to seal; from 7.4 to
    13.9 for seal to bear, and 49.2 for fish to bear. For individual PCB
    homologues, for instance, for fish to bear, these factors ranged from
    <0.5 (tetrachlorinated PCBs) to 263.4 for heptachlorinated PCBs.

    Niimi & Oliver (1989b) monitored the presence of 92 monochloro- to
    decachlorobiphenyl congeners in brown and lake trout, small and large
    rainbow trout, and small and large coho salmon from Lake Ontario. Each
    sample consisted of 8-12 fish. The highest concentrations were among
    the penta- and hexachlorobiphenyl homologues, with 2,4,5,2',4',5'-
    hexachlorobiphenyl the most common congener.

    Total congener concentrations ranged from 1 to 10 mg/kg in whole fish
    and from 0.3 to 4 mg/kg in muscle. The 10 most common PCB isomers were
    84, 87/97, 101, 110, 118, 138, 149, 153, and 180, and represented 52%
    of the total content. This value did not appear to be influenced by
    species or by total concentration.

    Huckins et al. (1988) collected fish (1-6 fish of 7 species) from the
    Waukegan Harbour of Lake Michigan, Illinois in 1978. The fish samples
    were analysed for the presence of 3,4,3',4'-tetrachloro- and
    2,3,4,3',4'-pentachlorobiphenyl. Total PCB congener residues averaged
    33.4 (2.4-56.6) mg/kg. The concentrations of 3,4,3',4'-tetra-
    chlorobiphenyl averaged 45.3 g/kg (2-89 g/kg) in the whole body. The
    concentrations for 2,3,4,3',4'-pentachlorobiphenyl averaged 229 g/kg
    (80-483 g/kg).

    Five times as much PCBs were found in herrings caught in
    industrialized areas of Sweden (near Stockholm) compared with those
    caught in the cleaner waters off the Swedish west coast. Levels in
    plankton fell progressively with increasing distance from
    industrialized areas (Jensen et al., 1972a).

    Holden (1973) found levels of up to 235 mg/kg in the blubber of seals
    sampled in the polluted coastal areas of the United Kingdom compared
    with lower levels (2 mg/kg) from unpolluted areas. Higher levels, (up
    to 88 mg/kg) were found in the blubber of toothed whales sampled in
    the North Sea, but none was detectable in similar species sampled off
    New Zealand and Surinam (Koeman et al., 1972).

    Peakall (1975) mapped out the global distribution of PCB levels in
    marine plankton. The values for the open North Atlantic (300-450 mg/kg
    lipid) were found to be very similar to those collected from polluted
    areas, such as the Baltic sea and the Firth of Clyde, in the United
    Kingdom. Values in the South Atlantic (12-64 mg/kg) were considerably
    lower. The highest values shown were for the Eastern coast of the USA
    (up to 3050 mg/kg). There were no values for the Pacific Ocean.

    When monitoring PCB levels in fish from the Mediterranean, Albaiges et
    al. (1987) found that territorial species reflected local inputs of
    the pollutant, but migratory species had baseline levels.

    Risebrough & de Lappe (1972) reported PCB levels higher than 3 mg/kg
    in fish from the industrialized areas of Tokyo Bay and New York Sound.

    Tanabe et al. (1986a) analysed Antarctic minke whales and found that
    they contained lower PCB levels than those caught in the Northern
    hemisphere (Tanabe et al., 1983). McClurg (1984) also found low levels
    of PCB in the Antarctic; Ross seals contained 0.09 mg/kg (in blubber).
    Mean levels of 0.69 mg PCB/kg (wet weight), found by Smillie & Waid
    (1987) in Australian fur seal blubber, were much lower than levels
    found in seals from the temperate Northern hemisphere. Similarly,
    Antarctic fish had very low PCB residues, ranging from 0.08 to
    0.77 g/kg wet weight (Subramanian et al., 1983).

    PCB residues in biota are usually highest near industrial sources, but
    this geographical distribution is becoming less pronounced. In fact,
    O'Shea et al. (1980) and Tanabe et al. (1988) found PCB levels in
    small oceanic cetaceans to be higher than those reported for
    terrestrial mammals and birds. For example, Tanabe et al. (1988) found
    the mean level of PCBs in the fatty tissue of the striped dolphin to
    be 36 mg/kg wet weight.

    Subramanian et al. (1986) analysed subcutaneous fat from Adelie
    penguins from the Antarctic and found PCB levels of 0.05 mg/kg fat
    weight. This is a factor of 100 lower than that in auks caught in the
    northern North Pacific (Tanaka & Ogi, 1984) and a factor of 10 000
    lower than residues found in the pectoral muscle (on a lipid weight
    basis) of herring gulls in the Baltic (Lemmetyinen et al., 1982).  Effect of dredging-contaminated sediment on organisms

    Dredging to remove contaminated sediments from the Shiawassee River,
    Michigan, increased the availability of PCBs, and, thus, residue
    levels, in freshwater clams (64.5-88 mg/kg dry weight) and in fish
    (fathead minnow; 13.8-18.3 mg/kg), both during dredging and up to 6
    months afterwards (Rice & White, 1987).  Relationship to lipid content of organisms

    PCBs are accumulated in lipid-rich tissues and care must be taken when
    interpreting results between species with different amounts of body
    fat. Jensen et al. (1969) found that PCB levels in herring and cod,
    from the same area of the Baltic Sea, were 0.27 and 0.033 mg/kg, on a
    wet weight basis, respectively, even though the cod is at a higher
    trophic level. The 2 species were found to have body fat contents of
    4.4 and 0.32%, respectively, and when the PCB residues were
    recalculated on a lipid weight basis, herring contained 6.8 mg/kg and
    cod, 11 mg/kg.

    PCBs are particularly accumulated in animals with large amounts of
    fat, such as seals, dolphins, porpoises, and whales (Tanabe, 1988) and
    in Arctic and Antarctic birds and mammals. Subramanian et al. (1986)
    found PCBs in all Adelie penguins sampled in the Antarctic, an area
    known to be relatively low in PCBs; the PCBs were mainly concentrated
    in fat-rich tissues. Kawai et al. (1988) measured PCBs in striped
    dolphins and found that the tissue level of PCBs depended entirely on
    their lipid content and, especially, on the amount of triglycerides in

    Redistribution of PCBs, from fat to other tissues, occurs in animals
    during periods of enforced starvation, such as seasonal food shortage,
    hibernation, migration, incubation, and the feeding of offspring.
    Subramanian et al. (1986) found that, as individuals Adelie penguins
    starved during incubation, residues of PCBs increased with declining
    fat reserves concomitant with tissue redistribution. Llorente et al.
    (1987) found that migratory duck species had a smaller percentage of
    the body burden of PCBs in adipose tissue than a resident species. A
    similar redistribution during starvation has been shown in the
    laboratory in European robins (Sdergren & Ulfstrand, 1972) and big
    brown bats (Clark & Prouty, 1977) (see sections and  Residues in different trophic levels and effects of diets

    In a study by Shaw & Connell (1982), bioaccumulation was increasingly
    evident in upper trophic level organisms, such as gulls and pelicans,
    in an Australian estuary compared with organisms from lower trophic
    levels. Veith et al. (1977) found typical PCB concentrations in Lake
    Superior biota to be 0.1 mg/kg for large zooplankton, 0.3 mg/kg for
    bottom fish, such as sculpins, and 1 mg/kg for pelagic fish.

    When various insects were sampled for PCB residues (Morse et al.,
    1987), levels in honey bees ranged from <0.1 to 1.5 mg/kg dry weight.
    PCB residues in other species ranged from <0.1 to 2.6 mg/kg, with
    predatory wasps containing the highest residues.

    Prestt et al. (1970) analysed the livers from various bird species in
    the United Kingdom. The highest PCB residues were found in freshwater,
    fish-eating species (up to approximately 900 mg/kg). The authors did
    not find any geographical pattern of distribution of PCBs in the
    species studied.

    Frank et al. (1975) collected birds' eggs from the Niagara peninsula
    in 1971. Eggs from carnivorous species of birds at the top of the
    aquatic food chain contained the highest levels of PCBs
    (3.5- 74 mg/kg). Terrestrial carnivores contained lower, but still
    relatively high, residues (0.2-1 mg/kg). Eggs from herbivorous and
    insectivorous birds contained much lower residues of PCBs. Again, eggs

    from terrestrial birds tended to contain lower levels (0.05-2 mg/kg)
    than those feeding on aquatic prey (0.14-4 mg/kg). Focardi et al.
    (1988) compared the PCB residues in the eggs of 8 species of water
    bird. The residues were found to be higher in fish-eating birds than
    in invertebrate feeders. The invertebrate feeders tended to contain
    higher percentages of the lower chlorinated congeners. Bird species
    that fed on other birds or fish had higher liver residues of PCBs than
    those feeding on mammals (Cooke et al., 1982). Peregrine falcons,
    herons, sparrowhawks, kingfishers, and great crested grebes had
    relatively high residues of PCBs. By contrast, golden eagles were only
    very lightly contaminated with PCBs.

    Bowes & Jonkel (1975) found a similar pattern in Arctic and subarctic
    food chains with PCB levels following the pattern: Arctic charfish
    < seals < adult polar bears < polar bear cubs.

    Mean PCB concentrations of 0.0018 mg/kg were found by Tanabe et al.
    (1984) in zooplankton, 0.048 mg/kg in myctophid, 0.068 mg/kg in squid,
    and 3.7 mg/kg in striped dolphin (all based on a whole-body, wet
    weight basis) sampled from the western North Pacific. The authors
    concluded that the bioaccumulation of chlorinated hydrocarbons was
    dependent on physical and chemical factors, such as water solubility
    and lipophilicity, in the lower trophic levels, whereas, in higher
    trophic levels, accumulation was affected by biochemical factors, such
    as the biodegradability of pollutants and the metabolizing capability
    of the organism.  Effects of age, sex, and reproductive status on uptake and

    Bache et al. (1972) found that the burden of PCBs increased with age
    in lake trout from Cayuga lake, Ithaca, New York, sampled in 1970
    (residues ranged from 0.6 to 30.4 mg PCBs/kg). An age- and
    length-related increase in PCBs was found in striped bass from the
    Hudson River and Long Island Sound; the author (Connell, 1987) stated
    that this observed relationship was due to the slow rate of
    bioaccumulation of the PCBs, particularly the higher chlorinated

    PCBs have been shown to accumulate with age in marine mammals, such as
    pinnipeds (Addison et al., 1973; Frank et al., 1973; Helle et al.,
    1983) and cetaceans (Gaskin et al., 1983; Aguilar & Borrell 1988;
    Subramanian et al., 1988). Helle et al. (1983) found mean levels of
    5.1 mg PCBs/kg (in extractable fat of blubber) in newly-born ringed
    seal pups, 17.3 mg/kg in seals of 2-4 months of age, and 65.3 mg/kg in
    sexually mature adults (4-12 years). However, lower levels of PCBs
    have been found in females compared with males (Martineau et al.,
    1987) and the age-related increase has often not been found in females

    (Addison & Smith, 1974). In many studies, while levels of PCBs in
    males have increased with age, those measured in females have fallen
    (Born et al., 1981; Gaskin et al., 1983; Aguilar & Borrell, 1988).
    Gaskin et al. (1983) found that PCB levels in the blubber of male
    harbour porpoises increased from 48.4 mg/kg at birth to 161 mg/kg
    after 8 years, whereas, in females, levels fell from 51 to 14.7 mg/kg.
    A significant decrease in the PCB levels was found by Subramanian et
    al. (1988) in female Dall's porpoises from 2 years of age onwards; 2
    years is required for the animals to reach sexual maturity. Excretion
    of PCBs during reproduction is known, from the laboratory, to be an
    important means of females losing residues. This PCB loss has been
    shown to be because of the transfer of PCBs to offspring via milk
    during lactation (Addison & Brodie, 1977). Addison & Brodie (1977)
    calculated that female grey seals excreted about 15% of their body
    burden of PCBs via lactation. In striped dolphins, females transferred
    between 72 and 98% of their body burden to the offspring (Fukushima &
    Kawai, 1981; Tanabe et al., 1982). It was suggested by Tanabe (1988)
    that such large transfer was because of the very high lipid content of
    the milk. Relocation of the PCB burden during pregnancy is generally
    thought not to be as important; in grey seals, the mother transfers
    only about 1% of her body burden to her offspring (Donkin et al.,
    1981) and in striped dolphins, only 4-9% (Fukushima & Kawai, 1981;
    Tanabe et al., 1982). However, Duinker & Hillebrand (1979) suggested
    that a much bigger percentage of female body burden (up to 15%) could
    be transferred to the fetus across the placenta of Harbour porpoise.

    Clark & Lamont (1976) calculated that female big brown bats
    transferred between 17 and 32% of their body burden of PCBs to their
    young, during gestation. The concentration of PCBs in adult females
    plus their litters declined with increasing age of the female. PCB
    levels were 0.83-3.6 mg Aroclor 1260/kg (wet weight) in adults and
    0.22-3.3 mg/kg in litters.

    When Passino & Kramer (1980) measured PCBs in deepwater ciscoes from
    Lake Superior, male fish contained significantly higher levels of PCBs
    (2.3 mg/kg wet weight) than females (1.2 mg/kg), eggs containing
    0.51 mg/kg. Lemmetyinen et al. (1982) found annual rates of
    elimination via egg production of 45% in the female Arctic tern and
    24% in the herring gull. Adelie penguins eliminated only 4% of their
    PCB body burden after laying their annual clutch of 2 eggs (Tanabe et
    al., 1986b). Elimination was thought to be dependent on the relative
    weights of the egg and mother.  Time trends in residues

    Buckley (1983) analysed various species of terrestrial plants from New
    York state. Total decreases of 42% in PCB residues were found between
    1978 and 1980.

    PCB levels in fish in the Hudson River, New York declined between 1977
    and 1981. The PCB levels were much higher in the Upper Hudson River
    (4217-1431 mg/kg of lipid), near to a major discharge of PCBs, than in
    the Lower Hudson River (1604-319 mg/kg) (Sloan et al., 1983).

    Frank et al. (1978) measured PCB levels in various fish species from
    Lakes Huron and Superior during the period 1968-76. PCB residues
    declined in lake trout and lake whitefish in Lake Superior between
    1971 and 1975, but increased slightly over the same period in bloaters
    and white sucker. In Lake Huron, PCB levels decreased between 1968 and
    1971, and, in alewife, rainbow smelt, and walleye, between 1975 and
    1976. In some of the study areas, residues increased in cisco, yellow
    perch, coho salmon, and splake but, at most locations, and, for other
    species analysed, no trends in PCB levels were found. St Amant et al.
    (1984) analysed fish from Lake Michigan between 1971 and 1981. An
    overall decrease in PCB levels was found for all species monitored
    except the walleye. Levels decreased from a maximum of 22.4 mg/kg at
    the beginning of the study to 3.8 mg/kg or less in 1981.

    Fish from all over the USA were analysed in 1980-81 by Schmitt et al.
    (1985) who found a significant downward trend (0.88-0.53 mg/kg PCB;
    wet weight) when mean residues were compared with fish collected
    between 1976 and 1977 (Schmitt et al., 1983). A similar downward
    pattern in residues was found in the Baltic when Moilanen et al.
    (1982) compared residues found in pike and herring caught between 1978
    and 1982 with those in fish sampled between 1972 and 1978 (Paasivirta
    & Linko, 1980). Haahti & Perttila (1988) found a continued decline in
    PCB residues between 1979 and 1986, when residues in herring muscle
    tissue decreased from 2.7-3.7 mg/kg to 0.3-1.1 mg/kg.

    An overall fall in PCB levels was found by Newton & Bogan (1978) in
    sparrowhawk eggs during the period 1971-74. Cooke et al. (1982)
    analysed liver samples from grey herons, kestrels, and barn owls for
    PCB residues during the period 1967-77. They found a significant
    decline in PCB residues over the sampling period in all 3 species. The
    mean residues in heron, kestrel, and barn owl for the period 1967-71
    were 5.77, 1.57, and 0.44 mg/kg, respectively, and for 1977, 0.56,
    0.6, and 0.15 mg/kg, respectively. However, Newton et al. (1986), when
    analysing sparrowhawk eggs from 1971-80, found that, although levels
    had fallen in the early 1970s, they had risen again in the late 1970s
    (mean PCB residues in eggs ranged from 16 to 293 mg/kg in lipid). Data
    on PCB residues in the livers of kestrel, sparrowhawk, heron,
    kingfisher, and the great crested grebe, collected from the late 1960s
    up to 1987, were analysed statistically by Newton & Haas (1989). For
    the great crested grebe, a significant overall decline in PCB residues
    was found when comparing data from 1987 with that from the 1960s. For
    the other species, there was no significant difference. Spitzer et al.

    (1978) reported that there was no significant change in PCB levels in
    osprey eggs collected from the Connecticut-New York area during the
    period 1969-76. Similarly, Wiemeyer et al. (1987) did not find any
    change in the carcase levels of PCBs in ospreys from the Eastern
    United States when comparing the 1971-73 and 1975-82 periods. They did
    find that adults contained significantly higher concentrations of PCBs
    than immature ospreys.

    Blus et al. (1979) analysed brown pelican eggs from South Carolina and
    Florida between 1969 and 1976. The highest levels of PCBs were found
    in South Carolina (means ranged from 5.25 to 7.63 mg/kg wet weight),
    but no significant trend was found during the study period. In
    Florida, the authors did not find any significant change in eggs
    collected from colonies in Florida Bay and on the Gulf Coast over the
    study period (means ranged from 0.62 to 1.18 mg/kg), but the Atlantic
    coastal colony showed a significant increase in PCB residues (from a
    mean of 2.68 to 6.12 mg/kg) between 1969 and 1976.

    In analysing herring gull eggs from the Great Lakes between 1974 and
    1978, Weseloh et al. (1979) found a significant decline in PCB
    residues from colonies on all the lakes. Lake Ontario, the most
    contaminated, showed the biggest decline from 170 to 75 mg PCBs/kg at
    one of the colonies, with other less contaminated Lakes, Huron,
    Superior, and Erie, showing levels in the range of 50-86 mg/kg in 1974
    and 32-46 mg/kg in 1978.

    Moksnes & Norheim (1986) analysed herring gull eggs collected from the
    Norwegian Coast between 1979 and 1981 and found that the PCB levels
    were not significantly different from those in eggs collected in 1969;
    mean PCB residues ranged from 1.2 to 6.7 mg/kg wet weight. They found
    a small but significant increase in the most persistent congeners and
    a significant decrease in DDE and the DDE/PCB ratio, but not in total
    PCB levels.

    An analysis of the eggs of double-crested cormorant (an inshore-
    subsurface feeder), Leach's storm petrel (an offshore-surface
    feeder) and Atlantic puffin (an offshore-subsurface feeder) was
    carried by Pearce et al. (1989), every 4 years, between 1968 and 1984.
    In the Bay of Fundy, Canada, PCB levels declined significantly during
    this period in all 3 species. PCB levels in the cormorant were
    consistently higher throughout than those in the other 2 species,
    ranging from 4 to 29.5 mg/kg (wet weight). Petrel and puffin eggs
    collected from the Atlantic Coast of Newfoundland showed lower levels
    than those in eggs from both the Bay of Fundy and the St Lawrence
    River estuary; as in the St Lawrence River, no significant trend in
    PCB levels was observed. A significant decline in PCB residues was
    found in gannet eggs collected during the same period from the gulf of
    St Lawrence (Elliott et al., 1988).

    The frequency of occurrence of measurable PCB residues has increased
    in large-scale sampling exercises; PCBs in mallard wings increased
    from 39% in 1976-77 (White, 1979) to 95% in 1979-80 (Cain, 1981). Cain
    & Bunck (1983) found that, in 1976, 21% of European starlings
    collected in the USA contained PCBs compared with 83% in 1979.

    Addison et al. (1986) analysed the blubber of Arctic ringed seals
     (Phoca hispida) from Holman Island, NWT, Canada, in 1981. They found
    mean PCB levels of 0.58 mg/kg (wet weight) in the females and
    1.28 mg/kg in the males. These concentrations were significantly lower
    than those detected in the same species from this area in 1972. Over
    this same period,  pp'-DDE levels, although at lower levels, also
    fell significantly, but it should be noted that total DDT levels in
    blubber are much lower than PCB levels and have not changed
    significantly.  Seasonal patterns in residues

    Jensen et al. (1969) observed that there was considerable seasonal
    variation in the fat content of herring caught in the Baltic Sea,
    ranging from 1% in the spring to 10% in the autumn and that this
    seasonal change in fat content led to seasonal changes in the tissue
    levels of PCBs.

    Cooke et al. (1982) found a seasonal pattern of PCB levels in European
    kestrels. Residues in both fat and liver were low in the autumn, but
    increased from about January, with a peak almost invariably occurring
    during the second quarter of the year (April, May, or June). Seasonal
    patterns were based on samples collected over a 10-year period.
    Similar trends were found in sparrowhawks and barn owls, but fewer
    samples were available.

    5.1.5  Appraisal

    PCB contamination is widespread and has been measured in a wide
    variety of biota between the 1960s and the present day. They are
    present throughout the world and, although initially concentrated in
    areas of high industrial activity, are now found in organisms living
    in remote areas, such as the oceans and the polar regions. In the
    past, PCB levels were positively correlated with areas of heavy
    industry and consequent discharge but, with the implementation of PCB
    controls, in some countries, these geographical differences are
    becoming less clear. Generally, levels of PCBs are declining in areas
    previously high in PCBs. However, time-trend analysis for the general
    environment shows little change in total PCBs since the late 1960s.
    The ratio of congeners is, as would be expected, changing, with lower
    chlorinated isomers disappearing and the more highly chlorinated ones
    becoming more dominant in environmental samples.

    PCBs are persistent and bioaccumulate in many organisms, because of
    their high lipid solubility and low biodegradability, and usually
    enter food-chains from water containing industrial discharge and by

    Because of their hydrophobic nature, PCBs are associated with both
    oildrop-like aggregates in the surface microlayer of water and with
    sediment on the bottom.

    They are accumulated by micro- and macroplankton organisms that live
    in the surface microlayer and by bottom-living organisms.

    5.2  Levels in animal feed

    The effects of pollution are seen in the use of fish-meal in poultry
    and fish farming. Kolbye (1972) sated that this may contain PCB levels
    of 0.6-4.5 mg/kg.

    Hansen et al. (1981) studied the transfer of PCBs in swine foraging on
    sewage sludge amended soils in 1975-76. Sixteen Berkshire sows were
    overwintered for 2 seasons on 4 experimental plots that had been
    treated with 0, 126, 252, or 504 tonnes/hectare (on a dry solids
    basis) of Chicago sewage sludge for the 8 preceding years. The
    estimated PCB residues in the soils of the 4 plots (average of 3-4
    samples) were 1.62, 1.88, 2.13, and 2.81 mg/kg dry weight (mean values
    of 3-4 samples/plot). The mean concentrations in fat of 3-4 sows per
    plot were 36  9, 106  64, 191  97 and 389  118 g/kg fat basis. Of
    the 12 individual congeners that were present in the fat, 3 accounted
    for more than 50% of the congeners, e.g., 2,3,4,2',4',5'-,
    2,4,5,2',4',5'-hexachlorobiphenyl and 2,3,4,5,2',4',5'-

    In vegetable animal feed (155 samples) originating from 5 areas of the
    world, samples, collected in 1984/85, contained PCB levels of 0.0009
    (Africa) up to 0.0093 mg/kg dry weight (Europe). In feed from North
    and South America and the Far-East, the levels were between 0.0024 and
    0.0066 mg/kg. Different types of feed originating from agriculture in
    the Federal Republic in Germany, collected in 1985, contained PCB
    levels of the order of 0.02 mg/kg dry weight. In feed (301 samples)
    originating from animals (exclusive fish meals), collected in 1985,
    0.021-0.036 mg/kg dry weight was found (DFG, 1988).

    Levels of 10-100 g/kg are given for groats, soybeans, and cotton
    seed, and a mean value of 18 g/kg is given for mixed feedstuffs. Fish
    meal contained levels of 110-330 g/kg (Klein, 1983).

    Samples of fish meal from different areas of the world, collected in
    1985, were analysed for the presence of PCBs. In 323 samples, the PCB
    contents varied between 0.006 and 0.055 mg/kg dry weight. The PCB
    congeners numbers 28, 138, and 153 were present in the highest
    quantities (DFG, 1988).

    Samples of fish meal from different areas of the world, collected in
    1985, were analysed for the presence of PCBs. In 323 samples, the PCB
    contents varied between 0.006 and 0.055 mg/kg dry weight. The PCB
    congeners numbers 28, 138, and 153 were present in the highest
    quantities (DFG, 1988).

    5.3  Levels in human food

    5.3.1  General

    Two general reviews of PCB residues in food, animal feed, human milk,
    plants, soils, and packaging materials have been published by Khan et
    al. (1976) and Sawhney & Hankin (1985).

    The PCB contents of a variety of foods on the Swedish market has been
    measured by West & Norn (1970a) and West et al. (1971). Less than
    0.1 mg/kg was found in samples of butter, margarine, vegetable oils,
    eggs, beef, lamb, chicken, bread, biscuits, and baby food; one sample
    of pork out of more than 100 had a PCB content of <0.5 mg/kg.

    In the period 1980-81, 5270 food samples were drawn at wholesale or
    production levels or at the site of importation including: butter,
    cheese, eggs, kidneys from pigs and cattle, and fat of poultry. Levels
    in Danish butter (99.4% of the samples) were below 0.05 mg/kg and
    those in imported butter (100%), below 0.125 mg/kg; Danish cheese
    (100% of the samples) levels were below 0.05 mg/kg and, in imported
    cheese, 82.4% of samples had levels below 0.125 mg/kg and the other
    17.6%, below 0.2 mg/kg; 100% of eggs had levels below 0.05 mg/kg, and
    100% of kidneys of pigs and cattle were below 0.15 mg/kg; 96% of
    poultry fat samples had levels below 0.15, and 4%, below 0.20 mg/kg,
    on a fat basis (not stated) (Statens Levnedsmiddelinstitut, Danmark,

    Mes et al. (1989b) studied the presence of specific isomers of PCB
    congeners in fatty foods of the Canadian diet. A total of 93 food
    composites from the cities of Ottawa and Halifax were analysed for 34
    PCB isomers, as part of a revised total diet programme. Each market
    basket comprised approximately 200 different food types collected from
    each of 4 major supermarkets in Ottawa during September 1985 and
    January 1986, and, in Halifax, in September 1986. Foods were used
     per se, or prepared and cooked in a manner ready for consumption,
    then composited to give 112 composites from each market basket.
    Thirty-one selected composites, representing the fatty foods were
    analysed from each market basket.

    PCB isomers 118, 138, 153, and 180 were found in all dairy products,
    except skimmed milk. Cheese and butter contained the highest levels of
    PCB residues. The residue level of isomer 118 (2,4,5,3',4'-
    pentachlorobiphenyl) in butter was the highest e.g., 0.7 g/kg, of all
    PCB isomers found in dairy products. Almost all meat, fish, and
    poultry contained PCB isomers 183 and 187. Occasionally, isomers 49,
    87, 185, and 189 were also present, but isomer 105 (2,3,4,3'4'-
    pentachlorobiphenyl), present in most dairy products, was only found
    in some beef samples. Fresh water fish contained most PCB isomers (28
    out of 34 selected PCB isomers), at levels considerably higher than
    those in any other meat, fish, or poultry samples. The level of isomer
    110 in fresh water fish was 3.05 g/kg. PCB isomers 138, 153, 180, and
    187 were present in almost all samples of meat and fish products,
    fats, oils, and soups. Cooking fats, salad oils, and margarine
    contained relatively low levels of PCB residues. PCB isomers 37, 49,
    87, 105, and 185 were not detected in meat and fish products, fats,
    oils, or soups.

    The calculated sum of all PCB isomer residues found in selected food
    commodities (except fish) ranged from 0.03 to 1.98 g/kg on a wet
    basis, and from 0.07 to 10.71 g/kg on a lipid basis, with mean values
    of 0.60 and 3.91 g/kg, respectively. However, the mean residue levels
    of fish and fish products were considerably higher, i.e., 10 and
    194 g/kg on a wet and lipid basis, respectively.

    The major PCB isomers in fatty foods were isomers 37, 52, 99, 110,
    118, 138, 153, 180, and 187.

    The PCB levels obtained in an extensive study by the US Food and Drug
    Administration are shown in Table 15. These values are considerably
    higher than those reported from Sweden, but they are probably biased,
    as they include samples originating from areas previously suspected of
    having been subject to local pollution.

    In a Canadian survey, PCB levels of less than 0.01 mg/kg were found in
    eggs (Mes et al., 1974) and a mean of 0.042 mg/kg was found in
    domestic and imported cheese with a maximum of 0.27 mg/kg (Villeneuve
    et al., 1973b).

    A preliminary study was carried out to estimate the dietary intake of
    PCBs in fresh food composites grown in Ontario in 1985. The following
    5 food composites: fresh meat and eggs, root vegetables (including
    potatoes), fresh fruit, leafy and other above-ground vegetables, and
    cow's milk were analysed. The concentrations in the different food
    composites were below 0.0005 mg/kg. The annual dietary intake of PCBs
    was estimated to be 32.6 g (Davies, 1988).

    In Japan, a similar range of PCB contents has been reported for most
    foods; however, some high levels have been reported for rice and
    vegetables harvested in fields polluted with PCBs (Environmental
    Sanitation Bureau, 1973). The PCB content of most fish on the market
    was less than 3 mg/kg.

    Table 15.  PCB levels in food in the USAa

    Food         % Positive      Level in positive samples (mg/kg)
                 (0.1 mg/kg)                                        
                                 Mean               Maximum

    Cheese            6          0.25                 1.0
    Milk              7          2.3                 27.8
    Eggs             29          0.55                 3.7
    Fish             54          1.87                35.3

    a  From: Kolbye (1972).

    Cantoni et al. (1988) analysed different food items, in 1985-87, in
    Italy, taking 20-60 samples per item. Different types of meat were
    analysed and the median concentrations were 0.25-0.50 mg/kg, on a fat
    basis. Twenty to 50% of the samples were positive. Poultry contained
    0.028 mg/kg, cow's milk 0.05 mg/kg, cream 0.027 mg/kg, butter
    0.065 mg/kg and fish 1.105 mg/kg, on a fat basis; 71% of fish samples
    contained PCBs.

    When the fat of poultry (42 samples) and 44 eggs was analysed, PCB
    values were below 0.3 mg/kg (Dutch Agricultural Advisory Commission,

    In the Federal Republic of Germany, wheat was analysed during the
    period 1972-82. The mean concentrations for 1972-78 ranged from 10 to
    30 g/kg; in the period 1980-82, the range was < 2.0-18 g/kg (Klein,
    1983). In wheat and rye (total 850 samples), median levels of
    0.4-1 g/kg product were found in 1984 (Codex Alimentarius, 1986). The
    concentrations found in other food items are summarized in Table 16.

    Samples of canned ham exported from Czechoslovakia to the USA in 1983
    contained PCBs levels of up to 4.8 mg/kg (Anon., 1983a,b).

        Table 16.  PCBs in food (1982) in the Federal Republic of Germanya

    Food               Total no.    Number of     Variation     Mean
                       of           samples       min-max       (g/kg)
                       samples      below         (g/kg)

    Milk                 854          234         < 2-3000       126.7 (FB)
    Beef                  76           43         < 10-687        72.4 (FB)
    Pork                  58           36         < 10-458        58.1 (FB)
    Poultry               64           61         < 10-85          7.3 (FB)c
    Meat products        185           86         < 4-2700       114.2 (FB)
    Eggs                  82           67         < 5-230          9.1 (FW)
    Fish (only            70            -           40-87         41.1 (FW)
    cod, herring,

    Food of plant origin

    Oil                  167          139         < 5-65           7.1 (FB)
    Cereals              345           44         < 2-30           6.7 (FW)
    Potatoes             106          106         < 2              -

    a  From: Klein (1983).
    b  Not stated.
    c  Only 3 samples.
       FB = fat basis
       FW = fresh weight.

    5.3.2  Drinking-water

    Ruoff et al. (1988) examined 83 drinking-water samples from the
    Federal Republic of Germany and from 5 other European countries for
    their contents of the PCB congeners 28, 52, 101, 138, 153, and 180.
    The average total content of the 6 congeners was 0.002 g/litre water.
    The average concentrations of the above-mentioned PCB congeners in the
    drinking-water of 6 countries were 0.0001, 0.001, 0.00018, 0.00035,
    0.00037, and 0.00042 g/litre. The variation between the 6 countries
    was quite small.

    The highest concentration of PCBs reported in domestic tap water was
    0.1 g/litre in the Kyoto area of Japan (Panel on Hazardous Trace
    Substances, 1972 cf. WHO/EURO, 1988), but, levels, more likely to be
    encountered, should not exceed 0.001 g/litre.

    In the FAO/WHO collaborating centres for the food contamination
    monitoring programme, the median levels were:


    Cereals                                  below  10 g/kg
    Vegetable fat/oils                       below  5 g/kg
    Fresh fruit and vegetables                      0.5-5 g/kg
    Animal fat (depending on type of
      animal and origin)                            20-240 g/kg
    Whole fluid cow's milk (depending
      on country)                                   10-200 g/kg
                                                    (on fat basis)
    Butter                                          30-80 g/kg
    Whole dried cow's milk                          20-50 g/kg
    Hen eggs                                        < 10 g/kg
    Fresh finfish                                   10-200 g/kg

    (WHO, 1985b).

    The contamination of a drinking-water system in Pickens County, South
    Carolina by PCBs discharged from a manufacturing facility was
    described by Billings et al. (1978). They observed that PCBs
    discharged by a capacitor manufacturing plant resulted in levels as
    high as 0.818 g/litre in finished potable water.

    5.3.3  Dairy products

    A number of data on food-producing animals have recently become
    available within the framework of the Joint FAO/WHO Food Contamination
    Monitoring Programme (JFCMP, 1985). All reported median values of PCBs
    in animal fat (excluding milk fat) were below the respective limits of
    detection, which varied from 0.001 mg/kg in the United Kingdom to a
    high of 0.5 mg/kg in Thailand and the USA. Data on PCB levels in cow's
    milk fat were supplied by the Federal Republic of Germany, Japan, the
    Netherlands, the United Kingdom, and the USA. The United Kingdom and
    the USA reported that median concentrations in cow's milk were below
    the detection limits of 0.5 g/kg and 0.5 mg/kg, respectively.

    The available data are summarized in Table 17.

    From the end of 1982 to the beginning of 1983, high levels of PCBs
    were detected in milk from several dairy farms in Switzerland. The
    investigations showed that the silo coatings and consequently the
    silage from the silos were the origin of the contamination of the
    milk. The PCB levels were between 0.80 and 3.80 mg/kg fat. PCB
    dissolution in acid juice, mechanical erosion of the coatings, and
    volatilization of the coating surface seemed to be the principal
    mechanisms explaining the migration of PCBs into the silage
    (Alencastro et al., 1984).

    Forty-two samples of cow's milk (14 samples in 1976, 14 in 1983, and
    14 in 1986) and 41 samples of market milk (10 in 1976, 16 in 1983, and
    15 in 1986) were analysed for PCBs, in Israel. During this period, a
    change was observed in the PCB distribution in the milk samples. The
    percentage of hexachlorobiphenyl decreased with time and the
    pentachlorobiphenyl increased (Pines et al., 1988).

    The monitoring data for dairy products from all over the world for
    1980-83 have been summarized by the Joint FAO/WHO Food Contamination
    Monitoring Programme (WHO, 1986a,b).

    5.3.4  Fish and shellfish

    A summary of the monitoring data on fish from all over the world for
    1980-83 has been published by the Joint FAO/WHO Food Contamination
    Monitoring Programme (WHO, 1986a,b).

    As might be expected, the PCB values found in fish depended on the fat
    content and the pollution of the fishing area (West & Norn, 1970a;
    Berglund, 1972).

    In a collaborative study by 7 national laboratories (International
    Council for the Exploration of the Sea, 1974), the PCB contents in the
    muscle tissue of fish taken from the North Sea were measured. A mean
    of 0.01 mg/kg was found in cod, while herring contained up to
    0.48 mg/kg, with most samples in the range of 0.1-0.2 mg/kg; plaice
    contained 0.1 mg/kg or less. Similar values were reported by Zitko
    (1974) for fish taken from the North Atlantic.

    Risebrough & de Lappe (1972) reported levels higher than 3 mg/kg in
    fish from New York Sound and Tokyo Bay, both very polluted areas. Even
    higher levels of PCBs were found in fish from polluted lakes and
    inland waterways, a level of 20 mg/kg being found in fish from Lake
    Ontario, and levels over 200 mg/kg in fish from the Hudson River
    (Stalling & Mayer, 1972). Similar correlations between pollution and
    PCB levels have been reported from the United Kingdom in fish
    (Portmann, 1970), and in mussels (Holdgate, 1971).

        Table 17.  Occurrence of PCBs in dairy products

    Country             Year          Product               Number of           Mean concentrations       Reference
                                                            samples             in mg/kg on fat basis

    North America

    USA                 1973-1974     milk (bulk)           198 (9 positive)      1.91 (0.32-4.99)        Willett (1980)


    Germany             1982-1986     milk                  3279                  0.09-0.14a              DFG (1988)
    (3 areas)           1983-1986     butter/cheese         2088                  0.05-0.11
    Westphalian area    1972-1974     butter                -                     0.38 (0.25-0.54)        Claus & Acker (1975)
    Northern part       1978-1980     milk                  -                     0.17-0.20               Codex Alimentarius
                        1984          milk                  3510                  0.013                   (1986)
    Northern part       -             butter                1836                  0.0077c                 Codex Alimentarius
                        -             meat and fat          957 (about 3/4        0.01b                   DFG (1988)
                                      cows entrails         51                    0.149b

    Sweden              1972-1977     beef, pork and meat   232 (217            < 0.001-0.01              Vaz et al. (1982)
                                      products (domestic    negative)           (whole product)
                                      and imported)

    Denmark             1981-1982     milk                  -                     0.10-0.13               Jensen (1983b)

    Table 17.  (cont'd).

    Country             Year          Product               Number of           Mean concentrations       Reference
                                                            samples             in mg/kg on fat basis

    Netherlands         1975-1977     milk                  315                 0.16 (0.06-0.33)          Gezondheidsraad
                        1980-1983     milk                  -                   0.07-0.13                 (1985)
                        1978-1984     milk                  2319                < 0.1-0.2                 Olling (1984)
                        1977-1981     cattle fat            -                   0.11b (< 0.05-0.55)       Greve & Wegman
                                      pork                  -                   0.07 (< 0.05-0.66)        (1983)
                        1983          fat of cattle, pork,  40-45               < 0.03b                   Dutch Agric. Adv.
                                      calves                                                              Comm. (1983)
                                      sheep                 22                  < 0.03b

    Switzerland         -             milk                  6                   0.034-0.144               Rappe et al. (1987)
    (6 locations)

    a  Major congeners were Nos. 138 and 153.
    b  Median value.
    c  Arithmetic mean.

    Jensen et al. (1969) found PCB levels of 0.27 mg/kg and 0.33 mg/kg,
    respectively, in the muscle tissue of herring and cod from the same
    area of the Baltic, though the cod is at a higher trophic stage. The 2
    species had 4.4 and 0.32% of extractable fat, respectively, and, when
    the PCB level was calculated on the fat content, values of 6.8 mg/kg
    for the herring and 11 mg/kg for the cod were obtained. Cod liver has
    a much higher fat content than cod muscle, and Jensen (1973) reported
    the ratio of PCB concentrations in cod liver and muscle to be over
    100, the maximum in liver being 59 mg/kg. Jensen et al. (1969)
    remarked that the considerable seasonal variation in the fat content
    of the herring, rising from 1% in spring to 10% in autumn, influenced
    the tissue level of PCBs.

    There are many examples of different PCB levels in similar species
    collected from areas of high and low pollution. Jensen et al. (1972b)
    found 5 times as much PCBs in herrings caught in waters off
    industrialized areas near Stockholm, as in herrings from the cleaner
    waters of the west coast of Sweden.

    Different freshwater and seawater fish were analysed for PCB contents,
    during the period 1981-83, in the Netherlands. Eel from different
    places over the period 1971-81 contained 0.2-13 mg/kg on a product
    basis (in the edible part). The median value was between 1 and
    2 mg/kg. Sea fish from the North Sea, such as herring and mackerel,
    contained 0.1-0.2 mg/kg, on a fat basis. The same level was found in
    shrimps and mussels (Freudenthal & Greve, 1973; Greve & Wegman, 1983;
    van der Kolk, personal communication, 1984a).

    The mean PCB contents in the liver of cod from the North Sea, North
    Atlantic, and Baltic Sea, were 2.1-5.7, 0.48, and 10.4-12.8 mg/kg,
    respectively (Klein, 1983).

    When fish from the North Atlantic, North Sea, and Baltic Sea, were
    collected in 1985, PCB concentrations of 0.098-0.123 mg/kg fillet
    weight were found in fish from the North Atlantic and North Sea and
    0.338 mg/kg fillet weight in fish from the Baltic Sea. In total, 60
    samples were analysed. The PCBs 101, 138, and 153 were the major
    congeners (DFG, 1988).

    The PCB concentration in freshwater fish of the River Rhine was found
    to be more than 2 mg/kg. The mean PCBs levels decreased, however, over
    the period 1976-81 from 1.92 to 0.38 mg/kg (fresh weight) (Klein,

    In 1984, PCB concentrations in freshwater fish (59 samples) collected
    in the River Rhine ranged from 0.742 to 1.017 mg/kg fillet weight. In
    this case, the major congeners were 138 and 153, but numbers 28, 52,
    101, 180 were also present. In total, 199 samples of eel were
    collected in a number of surface waters and analysed for the presence
    of PCBs. The levels ranged from 1.42 to 6.51 mg/kg fresh weight. In
    studies reported by DFG (1988), the highest levels of PCBs were found
    in the River Rhine.

    In the United Kingdom, fish and shellfish were analysed for PCBs
    during the period 1982-84 (HMSO, 1986). The results are summarized in
    Table 18.

        Table 18.  PCB levels in marine fish and shellfisha

    Year      Product                        Tissue      No. of      Range (mg/kg)

    1982      Marine fish (from England)     liver       381         0.3-4.1
              (7 types of fish)

    1982      Marine fish (from England)     muscle      326         0.03-0.13
              (7 types of fish)

    1983      Marine fish (imported)         muscle      102         nd-0.06
              (5 types of fish)

    1983      Shellfish (imported)           muscle      53          nd-0.06
              (4 types of shellfish)

    1984      Fish oils                                  16          0.11-2.3

    a  From: HMSO (1986).

    Different types of marine fish and shellfish from different areas in
    the United Kingdom were analysed during the period 1977-84. Those from
    the North Sea coast contained concentrations in the range of 0.04-5.7
    and < 0.001-0.058 mg/kg, respectively, while those from the English
    channel contained < 0.05-6.9 and < 0.006-0.1 mg/kg, respectively,
    and those from the West coast, < 0.002-8.4 and < 0.001-0.25 mg/kg
    wet weight. PCB concentrations in fish livers of 0.2 up to 12.9 mg/kg
    wet weight were found during this period (Franklin, 1987).

    When samples of fish of different species, collected from major USA
    watersheds in 1976, were analysed, PCBs were found in 93% of the
    samples. Fifty-eight of the samples had levels exceeding 5 mg/kg, on a
    whole fish basis. The PCB concentrations ranged from less than 0.3 to
    140 mg/kg, on a whole fish basis (Veith et al., 1979).

    Maack & Sonzogni (1988) analysed 98 fish (14 species) of different
    sizes from Wisconsin waters, for the presence of PCB congeners. Among
    the most prominent congeners were numbers 153/132, 138, 66/95, 110,
    180, 70/76, 146, 28/31, 149, 118, and 105. The total PCBs (determined
    by adding individual congener concentrations) ranged from 0.070 to
    7.0 mg/kg. The mean concentration was 1.3 mg/kg.

    Blue crabs ( Callinectes sapidus, an important member of the
    estuarine food web), collected from Campbell Creek and surroundings in
    South Carolina, were analysed for PCBs in 1985. The highest mean total
    concentration was 0.861 mg/kg muscle tissue. In 1986, the mean
    concentrations in blue crab collected by 8 stations in the same area
    ranged from 0.026 to 0.361 mg/kg muscle tissue. Blue crab (15 samples)
    collected from the coast of South Carolina, contained concentrations
    of < 0.020-0.372 mg/kg tissue (Marcus & Mathews, 1987).

    PCBs concentrations in sea fish were determined in 1971-77 in Japan.
    In-shore fish (90 samples) showed concentrations of 0.2-0.72 mg/kg
    fresh weight and pelagic fish (112 samples), 0.005-0.265 mg/kg fresh
    weight (Watanabe et al., 1979).

    Data on individual species of fish, submitted by Japan, showed the
    following median levels: barracuda, 70 g/kg; conger eel, 290 g/kg;
    croaker, 200 g/kg; flounder (yellow-tail), 90 g/kg; hair-tail,
    100 g/kg; mullet, 84 g/kg; and seabass, 110 g/kg. Median levels for
    other species of fish, such as cod, mackerel, pacific saury, rockfish,
    salmon, and sardines, were below 100 g/kg (WHO, 1986b).

    Using a very sensitive analytical method, Tanabe et al. (1987) found
    the toxic non- ortho-substituted coplanar 3,4,3',4'-tetrachloro-,
    3,4,5,3',4'-pentachloro-, and 3,4,5,3',4',5'-hexachlorobiphenyl in
    finless porpoise, at concentrations of 13.5, 0.89, and 0.64 g/kg,

    Blue mussel  (Mytilus edulis) was collected from coastal areas near
    Osaka and Hokkaido, Japan, in 1984-86. Depending on the site of
    collection, the average PCB concentrations (11-13 samples) ranged from
    0.56 to 65.0 g/kg (Miyata et al., 1987).

    5.3.5  Influence of food processing

    Fifty striped bass  (Morone saxatilis) were analysed for the presence
    of PCBs in the fish fillets before, and after, boiling, steaming,
    baking, frying, microwaving, or poaching, to study the possible
    reduction of the PCB residues by these cooking procedures. PCB
    contents were reduced by approximately 10%, by all 6 methods of
    cooking. No significant reductions were observed with the other
    cooking methods (Armbruster et al., 1987).

    5.3.6  Food contamination by packaging materials

    When Villeneuve et al. (1973a) analysed packaged food in Canada, they
    found that 66.7% of the samples contained PCB levels of less than
    0.01 mg/kg, 30.7% contained between 0.01 and 1 mg/kg, and 2.6%
    contained more than 1 mg/kg. The highest PCB levels were in a rice
    sample (2.1 mg/kg), where the packaging material contained 31 mg/kg,
    and in a dried fruit sample (4.5 mg/kg), in a container containing
    76 mg/kg. In a survey of packaging containers, approximately 80% were
    found to contain PCB levels of less than 1 mg/kg, while about 4%
    contained levels higher than 10 mg/kg. The most likely source of PCBs
    in packaging materials was the recycling of waste paper containing
    pressure-sensitive duplicating paper (carbonless copying paper)
    (Masuda et al., 1972).

    Relatively high PCB levels in some packaged foods in Sweden, mainly of
    imported origin, could be attributed to migration from the packaging
    material (West et al., 1971). The highest level encountered was
    11 mg/kg in a childrens' breakfast cereal; PCB levels of 70 mg/kg and
    700 mg/kg were found in the material of the inner bag containing this
    product and in the outer cardboard container, respectively. Up to
    2000 mg/kg was found in cartons of other samples.

    In the United Kingdom, levels in imported waste-paper, which could be
    contaminated with PCBs from carbonless copying paper and subsequently
    used to manufacture food contact paper and board materials, were found
    to be low, compared with the 10 mg/kg limit for PCBs recommended by
    the British Paper and Board Industry Federation for food contact
    materials (HMSO, 1989).

    5.3.7  Appraisal

    Foods have become contaminated with PCBs by 3 main routes:

    *   accumulation of PCBs in the different food-chains in the
        environment and consumption of fish, birds, or other animals and

    *   direct contamination of food or animal feed by an industrial

    *   migration from packaging materials into food.

    During the past years, many thousands of samples of different
    foodstuffs have been analysed for PCB contamination. The most common
    foodstuffs analysed have been fish, meat, and milk. Many fish samples
    have been taken in an effort to monitor aquatic pollution. In
    addition, samples have been taken, for regulatory or similar purposes,
    from sources suspected of being relatively highly contaminated. The
    fact that most samples have not been taken at random, jeopardizes the
    proper assessment of the exposure of the general population.

    5.4  General population exposure

    5.4.1  Air

    Relatively high levels of PCBs have been detected in indoor air,
    especially in kitchens and offices with electric installations
    (Jensen, 1983a) (section 3.2.4 and 5.1.1).

    Results from the US EPA indicate PCB concentrations in the air ranging
    from 1 up to 50 ng/m3; similar results have been reported from Japan
    (WHO, 1976). Assuming a level of 5 ng PCBs/m3 in urban air, a
    breathing rate of 22 m3/day, retention and absorption of inhaled
    particles/vapour of 50%, and a mean residence time of PCBs in the body
    of 3 years, air would contribute 0.8 g/kg to the PCB concentration in
    the body. Higher concentrations of PCBs in indoor air could increase
    this estimate (WHO/EURO, 1988).

    Van der Kolk (1985) calculated air intake through inhalation for the
    Dutch population of about 36 ng/day, a quantity approximately 1000
    times lower than the intake with food.

    During the manufacture, formulation, or use of PCBs, where levels in
    the workroom air correspond to exposure limit values, varying between
    0.1 mg/m3 and 1 mg/m3, the calculated mean intakes would range
    between 1 and 10 mg during an 8-h workshift. In some occupational
    situations, much higher concentrations have been measured and the
    estimates of intakes would be higher (WHO/EURO, 1987).

    5.4.2  Drinking-water

    Levels reported in drinking-water are typically between 0.1 and
    0.5 ng/litre. Even assuming a PCB level of 2 ng/litre in
    drinking-water, consumption of 2 litre/day contributes 0.04 g/kg body
    weight to the PCB concentration in the body. This additional quantity
    is negligible in comparison with the intake via food (WHO/EURO, 1988).

    5.4.3  Intake by infants through mother's milk

    The daily intake of PCBs was calculated in breast-fed infants in the
    countries participating in a monitoring study by Slorach & Vaz (1983,
    1985) (Table 19).

    The intakes in EEC countries were calculated to range from 3 to
    11 g/kg body weight per day, compared with 0.12-0.3 g/kg body weight
    for bottle-fed infants in Denmark (WHO/EURO, 1985).

    In Yusho infants with clinical symptoms of poisoning, the daily intake
    of PCBs with breast milk was calculated to be 70 g/kg body weight
    (Jensen, 1983b) (see section

    5.4.4  Infant and toddler total diet

    Johnson et al. (1979) analysed the average diet of 6-month-old infants
    and 2-year-old toddlers for the presence of PCBs. Ten market baskets
    were collected in 10 cities in the USA. The foods were prepared in the
    manner in which they would be prepared and served in the home. Trace
    amounts of PCBs were detected in only one infant and one toddler diet.

    In the USA, Gartrell et al. (1986b) found a daily intake of 0.011 g
    PCBs/kg body weight in infants consuming infant diets in 1978. In the
    years 1979, 1980, and 1981/82, the intake was below the detection
    level. The intake by toddlers was 0.099 g/kg body weight in 1978 and
    not detectable in the following 3 years.

    Tuinstra et al. (1985a) analysed samples of infant food from the Dutch
    market and found average PCB levels of 0.1-0.2 g/kg food (the maximum
    level found was 1.1 g/kg).

    5.4.5  Total intake by adults via food

    The oral consumption of contaminated products is presumed to be the
    main route of exposure to the PCBs.

        Table 19.  Calculated daily intakes of PCBs by breast-fed infants (g/kg body weight)a

    Country/area    Year(s)    Calculation according     Calculation according
                               to US FDA methodb         to national methodc

                               median        maximum     median        maximum

    Brussels        1982       3.6           10.4        NR            NR

    Beijing         1982       NR            NR          0.45d         0.45d

    Jerusalem       1981/82    2.0           9.5         NR            NR

    Hanau           1981       NR            NR          9.5           45

    Osaka           1980/81    1.6           4.4         2.3           6.3

    Uppsala         1981       4.4           8.1         5.9           11

    22 states       1979       4.5e          13.5        4.5e          22.5

    Zagreb          1981/82    2.8           7.2         2.8           7.7

    a  Assuming a milk consumption of ca 130 g/kg body weight and a milk fat content of
       3.5% (w/w). Calculations based on data for all mothers studied. Results for
       different methods of PCB analysis shown separately.
       From: Slorach & Vaz (1983, 1985); Van der Kolk (1984b).
    b  Sawyer method.
    c  "Own method".
    d  PCB level below limit of detection (0.1 mg/kg fat) in milk samples.
    e  PCB level below limit of detection (1 mg/kg fat) in milk samples.
    NR = No data on levels in milk reported.

    It has been stated that the major part of the human dietary intake of
    PCBs is from fish (Berglund, 1972; Hammond, 1972). This may well be
    true in areas such as Japan or certain localities near the North
    American Great Lakes, where fish from polluted waters may form a
    relatively large part of the diet. Several investigators from Japan
    have measured the daily intake of PCBs in food; the highest mean value
    recorded was 48 g/day, of which 90% was from fish (Kobayashi, 1972);
    the lowest was 8 g/day (Ushio et al., 1974).

    In much of Europe and North America, however, the daily intake of fish
    is in the region of 30-40 g, and most of the fish is taken from waters
    of low pollution with PCB levels in the fish not exceeding 0.1 mg/kg.
    Berglund (1972) has estimated that the daily intake of PCBs from fish
    in Sweden is in the region of 1 g, though if the fish consumed were
    solely Baltic herring, the intake would be about 10 g/person. It is
    difficult to make an assessment of the PCB intake from foods other
    than fish. West et al. (1971) in their extensive study of the
    Swedish diet, reported that most foods contained PCB levels of less
    than 0.1 mg/kg; and concluded that this corresponds to a daily intake
    of less than 100 g.

    Weekly intakes in the range of 23-889 g/person have been reported
    from the USA (OTA, 1979). The higher range concerns people consuming
    more than 12 kg/year of Lake Michigan fish.

    The intake of total PCBs by the general adult population depends
    greatly on the geographical area and food habits.

    5.4.6  Total diet/market-basket studies

    Data on total-diet studies of PCBs have been reported from a few
    countries. These reported intakes show a wide variation, which can
    partially be explained by methodological factors, such as the ways in
    which samples below the limits of determination are considered,
    especially when noting the different limits of determination.
    Considering the available data, an average intake of 5-15 g/day for
    the non-occupationally exposed population in industrialized countries
    may be the best available estimation.

    These estimates apply to the average diet of an average adult citizen.
    In practice, few people are really "average" in their consumption
    pattern. Given the widespread nature of the contamination, however, a
    higher intake in one food group is more or less balanced by a lower
    intake in another food group with an equal calorie intake. Total
    intake will certainly be higher for diets with a more than average
    calorie content (van der Kolk, 1985).

    Gartrell et al. (1985) determined the total intake of PCBs by 16- to
    19-year-old males in the USA. The samples represent a typical 14-day
    diet. Approximately 120 individual food items (of 12 food groups),
    including drinking-water, were collected for each market-basket sample
    in 20 cities in the period 1979-80. Only 2 samples of meat, fish, and
    poultry contained PCBs with an average concentration of 0.002 mg/kg.

    Gartrell et al. (1985, 1986a) reported a daily intake in the USA of
    0.016, 0.027, 0.014, 0.008, and 0.003 g PCBs/kg body weight during
    the years 1977, 1978, 1979, 1980, and 1981/82, respectively.

    Manske & Johnson (1975) collected 35 market baskets in 32 cities over
    the period 1971-72. PCB residues were found in the range of
    0.035-0.15 mg/kg in 51 composites. Fish and oils, fats, and
    shortenings contained the highest levels. The same authors (Manske &
    Johnson, 1977) carried out a market-basket study representing the
    basic 2-week diet of a 16- to 19-year-old male. The various foods were
    prepared in the manner in which they would normally be served and
    eaten. Thirty market-baskets, containing 12 classes of foods (in total
    360 composites) were collected in 30 cities in the period 1973-74. A
    trace of PCB was found once in whole milk, ground beef, and fish

    The FDA revised the concept of the Total Diet Study in 1982. As
    discussed by Gunderson (1988b), the Total Diet Study conducted before
    1982 was based on a "composite sample approach", regardless of the
    diet involved. The revised study is based on updated dietary survey
    information and allows the "total diet" of the US population to be
    represented by a relatively small number of food items for a greater
    number of age/sex groups. The daily intake expressed in ng/kg body
    weight per day for PCBs (Aroclor 1221, 1242, and 1254) in 1982-84 for
    the age groups 6-11 months, 2 years, 14-16-year females, 14-16-year
    males, 25-30-year females, 25-30-year males, 60-65-year females and
    60-65-year males were: 0.8, 1.2, 0.4, 0.5, 0.5, 0.6, 0.4, and
    0.5 ng/kg body weight per day, respectively (Gunderson, 1988b).

    Foods, representative of Canadian eating habits, as determined by a
    national nutritional survey, were prepared for eating, categorized,
    and blended into 11 different composites representing the dietary
    intake for 5 cities over the period 1976-78. It concerned 194 samples,
    collected in winter and in summer. The average dietary intake was
    0.001 g PCB/kg body weight (McLeod et al., 1980).

    Over a period of 2 years, 126 different food items of a market-basket
    of 16- to 18-year-old males were purchased every 2 months in the
    period 1976-78, in the Netherlands. The foodstuffs were prepared for
    eating and were combined in 12 commodity groups. The mean
    concentration and range of PCBs in 5 food classes was:


    Class                      Mean concentration      Range
                               (mg/kg on fat basis)

    Meat, poultry, and eggs         -                  0.13-0.17 (2)a
    Fish                            0.07               0.04-0.24 (7)
    Dairy products                  -                  0.04-0.06 (2)
    Sugar and sweets                -                  0.08 (1)
    Drinks, drinking-water          -                  0.035 (1)

    a  In parentheses: number of positive composites.

    The authors calculated a daily intake of PCBs of 15 g/person (a
    maximum level was 90 g/person (de Vos et al., 1984). In the period
    May-July 1976, 100 total diets (summer meals) were collected and
    besides organochlorine pesticides, PCBs were determined as
    decachlorobiphenyl, after perchloration, and calculated as Aroclor
    1260. The mean intake of PCBs/person per day was 11.6 g with a range
    of 3-71 g (Greve & van Hulst, 1977; Greve & Wegman, 1983; van der
    Kolk, 1985).

    In 1978, another survey was carried out with 100 total diets during
    the winter (winter meals). It was estimated that the daily intake was
    6 g/person (range 1-19 g).

    Zimmerli & Marek (1973) studied the total human intake of PCBs from
    prepared meals in 1971-72 in Bern, Switzerland. Five typical total
    diets were composed and analysed. The intake of PCBs, especially with
    daily diets containing cheese, meat, fish, or fat, ranged from 6 to
    84 g.

    According to a calculation by Summerman et al. (1978), the average
    weekly intake of PCBs in the Federal Republic of Germany was about 215
    and 268 g/week for females and males, respectively. Much lower
    figures, 36-44 g/week, were calculated by Klein (1983).

    A survey of the daily PCB intake from the total diet of Japanese women
    (number of samples varied from 18 to 60) was performed for the years
    1972-76. The daily intake of PCBs averaged approximately 10 g/person
    (range 2.8-21.2 g). The main source of PCBs in the diet of Japan was
    in-shore fish. There was no clear change in daily intake over the
    5-year period studied (Watanabe et al., 1979).

    Ushio & Doguchi (1977) studied the dietary intake of PCBs in Tokyo.
    They found an average daily intake of PCBs of 6.3 g/person (range,
    trace-17 g/person). It was concluded that the dietary daily intake of
    PCBs for the majority of the population of Tokyo rarely exceeded
    20 g/person, when no heavily contaminated fish were consumed.

    Yakushiji et al. (1977) found that the PCB daily intake through meals
    of unexposed adults living in Osaka prefecture, was 3-20 g/day.

    Data for PCBs in the diets of Canada, Guatamala, Japan, the United
    Kingdom, and the USA over the period 1972-83 were summarized by
    Gorchev & Jelinek (1985). The mean dietary intake reported was at, or
    below, 0.06 g/kg body weight, the mean intake per person ranged from
    < 0.01 to 0.12 g/kg body weight (Slorach et al., 1982; WHO, 1986b).

    5.4.7  Total intake of major congeners by adults via food

    In the Federal Republic of Germany, the daily intake of the 3 PCB
    congeners numbers 138, 153, and 180, together with the different food
    items, was calculated. The intake (g/day) with meat and meat products
    was 0.30; with fish and fish products 0.36; eggs and egg products
    0.008; milk and milk products 0.40; cheese 0.11; butter 0.39; fats and
    oil 0.098; bread and pastries 0.17; potatoes 0.081; vegetables 0.11
    and fruits 0.082 (DFG, 1988).

    5.4.8  Time trends in different matrices

    Although many countries introduced severe restrictions on the
    manufacture, use, and disposal of PCBs many years ago, it is difficult
    to discern any marked decline in the levels in human milk fat, from
    the published data.

    Levels of PCBs were estimated in 1085 samples of different cereals,
    collected in the Federal Republic of Germany over the period
    1972/74-1984. The levels, which were the highest in 1972/74 0.04 mg/kg
    (0.005-0.12 mg/kg), decreased during the years to 0.004-0.005 mg/kg
    dry weight in 1984 (DFG, 1988).

    Data from the Federal Republic of Germany showed no clear trend in PCB
    levels in human milk during 1975-79 (Slorach et al., 1982). The same
    was found in the Netherlands over the period 1974-83 (Greve & Wegman,

    Japanese data showed a decline in PCB levels in the fat of whole cow's
    milk during the period 1972-79. A decline was also found in PCB levels
    in finfish from coastal waters and in total marine fish (Slorach et
    al., 1982).

    A downward trend was found in human milk from Japan over the period
    1972-80. Each year, a large number of samples (361-877 samples/year)
    were analysed. In 1972, the median level was about 0.8 mg/kg and, in
    1980, 0.5 mg/kg, on a fat basis. A gradual decline was observed
    (Slorach & Vaz, 1983).

    In Canada, human milk and adipose tissue from Ontario residents were
    analysed over the period 1969-74. The values found did not indicate a
    trend in this period.

    The mean total PCB intakes determined in the FDA Total Diet Study, for
    the period 1971-87, for a typical "adult" diet, represented in Fig. 4,
    reflect that of a 14- to 16-year-old male during 1982-87. A clear
    decline was shown from approximately 7 g/person per day to less than
    0.1 g/person per day (Gunderson, 1988a).

    The daily intake of PCBs, expressed as ng/kg body weight per day, by
    6-month-old and 2-year-old children in the years 1980, 1981/82, and
    1982/84 did not show a trend, while, in adults, a decrease from 8 to
    0.5 ng/kg body weight per day was observed over the same years
    (Gunderson, 1988b).

    5.5  Concentrations in the body tissues of the general population

    The PCB levels in body tissues are a good indication of the overall
    and total exposure of the body to PCBs.

    Several factors may influence the concentrations of PCBs in body
    tissues, including duration and level of exposure, the route and
    pattern of exposure, the chemical structure of the PCB (degree and
    position of chlorination in the molecule), the amount of adipose
    tissue, other simultaneous exposures, as well as other biological

    5.5.1  Adipose tissue

    In general, while highly chlorinated congeners accumulate more easily,
    a lower degree of substitution provides more possibilities for
    hydroxylation and facilitates excretion. Factors other than the degree
    of substitution also affect accumulation, particularly the position
    and pattern of substitution (WHO/EURO, 1987).

    The available information on the occurrence of PCBs in the body fat of
    the general population is summarized in Table 20.

    FIGURE 4

        Table 20.  Concentrations of PCBs in the body fat of the general population

    Country                 Year            Number of samples   Mean concentration in mg/kg   Reference
                                                                on fat basis (range)

    North America

    USA (18 states)         -               637                 < 1    (68.9%)e               Yobs (1972)
                                                                < 1-2  (25.9%)e               Price & Welch (1972)
                                                                > 2    (5.2%)f

    Northeast Louisiana     1980            8                   1.04   (0.38-2.33)            Holt et al. (1986)
                            1984            10                  1.23   (0.65-1.44)

    Texas                   1969-1972       88 (15 positive)    1.7    (0.6-9.9)              Burns (1974)

    New York                -               101 (women)         3.4  1.1                     Bush et al. (1984)

    (urban and rural

    Canada                  -               99                  0.94   (0.04-6.8)a            Mes et al. (1982)

    Ontario                 1976 and        570                 2.1-2.2                       Frank et al. (1988)

    Table 20.  (cont'd).

    Country                 Year            Number of samples   Mean concentration in mg/kg   Reference
                                                                on fat basis (range)


    Japan (Kochi area)      -               -                   2.86   (maximum 7.5)          Nishimoto et al. (1972a,b)

    Japan                   1971-1982       -                   0.5-6.0a                      Katsunuma et al. (1985)

    Tokyo                   1974            30                  1.04   (0.38-2.5)             Fukano & Doguchi (1977)

    Japan                   -               241                 0.30-1.48                     Curley et al. (1973b)

    New Zealand             -               51                  0.82                          Solly & Shanks (1974)


    South Africa            1982            63                  0.15-5.18                     van Dijk et al. (1987)


    Austria (Vienna area)   -               32                  0.3-7.3                       Pesendorfer et al. (1973)

    Finland                 -               105                 0.2                           Mussalo-Rauhamaa et al.

    Germany, Federal        -               20                  5.7                           Acker & Schulte (1970)
    Republic of             -               282                 8.3                           Acker & Schulte (1974)
                            1982-1983       50b                 0.5-1.5                       Niessen et al. (1984)

    Italy (Siena)           1983-1984       26                  1.75c  (dry weight)           Focardi et al. (1986)

    Table 20.  (cont'd).

    Country                 Year            Number of samples   Mean concentration in mg/kg   Reference
                                                                on fat basis (range)

    Netherlands             1973-1983       24-78 per year      1.6-2.5d                      Greve & van Harten
                                                                                              Greve & Wegman (1983,

    Norway (Oslo)           -               40                  1.6                           Bjerk (1972)

    Spain                   1985-1987       14                  1.68                          Camps et al. (1989)

    United Kingdom          -               201                 < 1.0                         Abbott et al. (1972)
                            1976-1977       236                 0.7 (nd-10)                   HMSO (1986)
                            1982-1983       187                 0.9 (0.1-6.9)

    a  Wet weight.
    b  34 infants, 14 children, and 2 older children.
    c  About 60% included only five congeners: Nos. 118, 138, 153, 170, 180.
    d  Median.
    e  Percentage of samples.  PCBs in the fetus

    PCBs are also present in serum and all organs of the body in
    proportion to their fat content. PCBs pass more, or less (depending on
    structure and chlorination), through the placenta into the fetus.
    Since the fetus has little adipose tissue until 7 months of age, PCB
    concentrations may be higher in vital organs, such as the adrenal
    gland, but available data suggest somewhat lower levels in the brain
    (Masuda et al., 1978a; Kodama & Ota, 1980).

    Masuda et al. (1978a) found PCB levels of 270-960 g/kg fat in adipose
    tissue samples of fetuses beyond 7 months of gestation. Levels in the
    adipose tissue of adult females from the same geographical area ranged
    from 270 to 1360 g/kg fat. The mean concentrations were 470 g/kg for
    fetuses and 780 g/kg for adult females. However, since the ranges
    showed an overlap and the number of samples was small, it is not clear
    whether this represents a true difference.  Congeners in adipose tissue

    Wegman & Berkhoff (1986) investigated the presence of the different
    congeners in 24 human fat samples, collected in 1984. The following
    congeners were present at the highest levels: 2,4,4'-trichloro-,
    2,4,5,2',5'-pentachloro-, 2,4,5,3',4'-pentachloro-, 2,3,4,2',3',4'-
    hexachloro, 2,3,4,2',4',5'-hexachloro-, 2,4,5,2',4',5'-hexachloro-,
    2,3,4,5,2',4',5'-heptachloro, 2,3,4,5,2',3',4',5'-octachloro, and

    Focardi & Romei (1987) analysed 30 samples of adipose tissue,
    obtained from patients in Siena, Italy, in 1986, for the presence of
    19 PCB congeners. The results indicate that the mean PCB (as sum of
    the congeners) concentration was 1063 g/kg dry weight (range
    391-1918 mg/kg). The major constituents of the PCBs (about 60%) were
    the isomers 99, 138, 153, 170, and 180.

    Human adipose tissue was analysed for 3 non- ortho chlorine
    substituted coplanar congeners: 3,4,3',4'-tetrachloro-, 3,4,5,3',4'-
    pentachloro- and 3,4,5,3',4',5'-hexachlorobiphenyl (Kannan et al.,
    1988). Twelve samples, from 7 male and 5 female persons were obtained
    from hospitals. The average total PCB concentrations were 1.22 and
    1.02 mg/kg (wet weight basis), respectively. The concentrations of the
    3 congeners were 94-860, 120-730, and 36-200 ng/kg, on a wet weight
    basis, respectively.

    5.5.2  Blood of the general population

    Finklea et al. (1972) studied human plasma of different races of the
    population (723 volunteers with ages ranging up to 60 years) of urban
    and rural areas of South Carolina. The average concentration was
    5 g/litre (range 0-29 g/litre). No age effect was found, but ethnic

    differences and ethnic residence interactions were significant. Kreiss
    (1985) found mean serum concentrations in the non-occupationally
    exposed population in the USA, of between 4 and 8 g/litre, with 95%
    of the individuals having serum PCB concentrations of less than
    20 g/litre. More data are summarized in Table 21.

    Maternal blood and fetal cord blood were collected from volunteers
    from an urban and rural vicinity in upstate New York. Whole blood
    samples were taken from 101 women (26  4 years) entering maternity
    facilities. Maternal blood contained 3.4  1.1 g PCBs/kg and fetal
    cord blood contained 2.4  1.0 g/kg whole blood. The PCB congeners
    making up these totals were surprisingly few; 38% of the total residue
    in the maternal blood and 21% of the fetal cord blood comprised only 4
    components, 2,4,4'-trichlorobiphenyl, 2,4,5,2',4',5'-hexachloro-,
    2,3,4,2',4',5'-hexachloro-, and 2,3,5,6,2',3',6'-heptachlorobiphenyl.
    The congener 2,5,2',5'-tetrachlorobiphenyl crossed the placenta
    preferentially (Bush et al., 1984).

    The concentrations of PCBs were determined in blood samples from 120
    women hospitalized for miscarriages and 120 full-term pregnancy
    controls. The average PCB level was higher in women with miscarriages
    than in control women (8.65 g/litre and 6.89 g/litre, respectively,
    as Fenclor 54 and 14.81 and 14.90 g/litre, respectively, as
    decachlorobiphenyl). The reproductive history of each woman was
    assessed together with confounding variables and with environmental
    exposure and food intake. Food consumption did not indicate diet as
    the main source of PCB intake (Leoni et al., 1989).

    A cross section of the population of Michigan was studied following an
    accidental exposure in 1978. Five years after the accident, PCB and
    PBB residues were measured in adipose tissue and serum. Serum levels
    of PCB were measured in 1681 adults and 1462 children. Children (430)
    were found to have uniform levels throughout the state (mean
    concentration 4  2 g/litre). In adults, the serum PCB levels were
    higher in the area with highest PBB levels. The mean serum PCB level
    was 21 g/litre, compared with control levels for the rest of the
    state of 9 g/litre. No sex difference was found (Wolff et al.,

        Table 21.  Concentrations of PCBs in whole blood of the general population

    Country                  Year          Number of samples   Mean concentration in             Reference
                                                               g/litre (range)


    Ontario area             1975-1976     118                 18                         Frank et al. (1988)
    (patients suspected      1980-1981
    of being exposed         1984


                             -             -                   3.2                        Doguchi & Fukano (1975)

                             -             28 (women)          2.6                        Kuwabara et al. (1978)

    (Osaka area)             1976          16 (women)          2.8 (1.7-4.6)              Kuwabara et al. (1979)
                             1972-1977     -                   3-4                        Yakushiji et al. (1977)

    farmers                  1978-1983     -                   trace-21.4b                Katsunuma et al. (1985)

    Tokyo                    1973          27                  3.19 (2.2-5.1)             Fukano & Doguchi (1977)
                             1975          10                  2.59 (1.8-3.8)

    Table 21. (cont'd).

    Country                  Year          Number of samples   Mean concentration in             Reference
                                                               g/litre (range)

    Finland                  -                                 3.1-12                     Karppanen & Kolho (1973)

    Netherlands              -             34 (women)          4.5 (nd-11.6)              Blok et al. (1984)
                             31 (men)      4.8 (1.0-17.1)

                             1978          48-127              3.1e                       Greve & Wegman (1983,
                             1980          samples/year        3.5                        1984)
                             1981                              4.4
                             1982                              4.4

    North America

    South Carolina           1968          723                 5 (4.2-5.5)a               Finklea et al. (1972)
    (urban and rural

    Michigan                 1973          1100                56c                        Kreiss (1985)
    (areas of Lake           1979-1981                         17.2-23.6c

    Lake Michigan            1985          196                 5.5  3.7                  Schwartz et al. (1983)
    (high fish

    Table 21. (cont'd).

    Country                  Year          Number of samples   Mean concentration in             Reference
                                                               g/litre (range)

    Yugoslavia               1984-1986     10f                 155 (35-480)d              Jan & Tratnik (1988a)
    (residents around                      19g                 11 (6-18)
    River Krupa;                           4h                  5 (2-7)
    contamination by a
    plant using PCBs)

    a  Plasma.
    b  Serum.
    c  Geometric mean.
    d  Arithmetic mean.
    e  Median concentration.
    f  Living close to plant.
    g  Living 1-3 km from plant.
    h  Non-exposed other areas.

    Specific PCB isomer levels in the blood of 30 children, ages 2-5
    years, residing in an area of PCB-contaminated soil in Canada, were
    compared with those of 25 children in a non-contaminated area. The sum
    of individual PCB isomer levels in the exposed and non-exposed group
    were not significantly different, e.g., 0.54 g/litre (range
    0.22-0.99 g/litre) and 0.88 g/litre (range 0.28-2.30 g/litre). The
    major component in both groups was 2,4,5,2',4',5'-hexa-chlorobiphenyl
    (Mes, 1987).

    High levels of PCBs were found in the blood (up to 100 g/litre) in
    patients with severe weight loss (Hesselberg & Scherr, 1974). This was
    attributed to the release of PCBs from the mobilization of fat.

    Greve & van Harten (1983b) studied the relationship between the levels
    of PCBs in the adipose tissue and in the blood of the same persons. A
    total of 48 persons were involved in this study. A concentration
    factor (concentration in adipose tissue divided by concentration in
    blood) of 660 was found.

    5.5.3  Human milk

    Human milk is the major source of exposure for breast-fed infants. The
    amount of human milk secreted varies widely. The composition of the
    milk is related to the amount secreted, the stage of lactation, the
    timing of withdrawal (early or late in feeding) and to individual
    variations among lactating women. The individual variations depend on
    maternal age, health, social class, and diet. The concentration of
    PCBs depends primarily on the lipid concentration in milk. Wide
    variations in published results are caused by inaccuracies inherent in
    the analytical methods used for the quantification of lipids, and
    whether the milk sample is collected early or late during the feeding
    period. The fat content increases during emptying, and the fat content
    of milk from the 2 breasts may differ. According to a recent
    determination, the fat level in human milk averages 2.6-4.5%
    (WHO/EURO, 1988).

    Whether the differences in concentration in various countries are
    merely a function of the analytical methods used and the type of
    samples collected or whether true differences in body burden exist, is
    not clear at present. For instance, some countries have reported
    levels of PCBs in human milk fat ranging from nondetectable to
    14 mg/kg, while, in other countries, the highest levels found have
    been around 3 mg/kg. Because of these variations, calculating an
    average dose for nursing infants is difficult. The same difficulties
    exist when attempts are made to investigate trends over time
    (WHO/EURO, 1988).

    The results of the older studies have been obtained with a less
    sophisticated method using packed column GC. With this method only a
    dozen peaks can be separated. The quantitative results are reported as
    "total PCB values", though different techniques of quantification and
    different types of calculations were used.

    In contrast with the situation with many organochlorine insecticides,
    the levels of PCBs in human milk fat are higher in European countries,
    Japan, and the USA than in China (Slorach & Vaz, 1983, 1985), and are
    significant, particularly in the highly industrialized countries.
    Results from a large number of countries have been summarized by
    Jensen (1983a, 1985, 1987), Acker et al., (1984), Katsunuma et al.
    (1985) (especially Japanese data; period 1972-83); and WHO/EURO,
    (1987, 1988). The countries concerned are: Argentina, Austria,
    Belgium, Canada, Finland, France, Federal Republic of Germany (Klein,
    1983), German Democratic Republic, Israel, Japan, the Netherlands,
    Norway, Poland, Romania, South Africa, Sweden, Switzerland, Turkey,
    United Kingdom, USA, USSR, and Yugoslavia. The average levels of PCBs
    in human milk do not appear to differ very much between the
    industrialized countries and range between 0.5 and 2 mg/kg milk fat,
    except in Czechoslovakia, the Federal Republic of Germany, India,
    Denmark and Italy, where levels up to 3 mg/kg milk fat were found
    (Jensen, 1983b; Acker et al., 1984) (Table 22).

    The variation in residue levels in human milk during lactation was
    investigated in 5 women in the Federal Republic of Germany. Month-mix
    samples, composed of breast milk samples collected weekly, were
    analysed over a lactation period of between 5 and 9 months. The ages
    of the women ranged from 23 to 36 years. The PCB concentrations were
    between 0.61 and 2.20 mg/kg, on a fat basis. While the concentrations
    remained relatively constant, some fluctuations were seen but no trend
    was observed over the lactation period investigated (Fooken & Butte,

    Breast milk samples from 16 women in Canada were analysed for PCBs at
    8 intervals (7, 14, 28, 42, 56, 70, 84, and 98 days) during the
    lactation period. The average PCB concentrations in breast milk varied
    between 22.8 and 29.7 g/kg whole milk. No clear decrease or increase
    was observed. The average milk/blood ratio for PCBs was 23 and
    remained relatively constant during lactation (Mes et al., 1984).

    Wolff (1983) reported the half-life of PCBs (percentage chlorine not
    specified) in breast milk to be 5-8 months and found that the
    concentration of PCBs in breast milk was 4-10 times that in the
    maternal blood. Similar results were reported by Jacobson et al.

        Table 22.  Concentrations of PCBs in breast milk of the general population

    Region                          Year           Number of             Mean concentration in         Reference
    Country                                        samples               mg/kg on fat basis (range)

    North America

    USA (Michigan)                  1977-1978      1057                  1.5 (maximum 5.1)             Wickizer et al. (1981);
                                                                                                       Wickizer & Brilliant (1981)

    Canada (Quebec)                 -              154                   0.84 (nd-4.34)                Dillon et al. (1981)

    Ontario                         1971-1974      -                     1.2 (0.1-3.0)                 Atkinson (1979)
                                    1978           215                   0.6  0.3

    Ontario                         1975-1985      348                   0.023 (0.016-0.033)a          Frank et al. (1988)

    Five regions across Canada      1982           210                   0.697                         Mes et al. (1986)

    Regina, Saskatchewan            1979           80                    0.0052 (0.001-0.019)a         Qureshi & Robertson (1987)

    Table 22.  (cont'd).

    Region                          Year           Number of             Mean concentration in         Reference
    Country                                        samples               mg/kg on fat basis (range)


    Japan (Osaka)                   1972-1977      -                     0.030-0.040a                  Yakushiji et al. (1977)
                                    1969-1976      19-52 each year       1-2                           Yakushiji et al. (1979)

    India (Ahmedabad)               1981-1982      50                    not present                   Jani et al. (1988)

    Hawaii (different islands)      1979-1980      54                    0.80  0.43 (0.13-2.2)        Takei et al. (1983)


    Germany,                        since 1970     several thousands     1.0-2.5 (98% of samples       Acker et al. (1984);
    Federal Republic of                                                  between 0.001-7.2)            Cetinkaya et al. (1984);
                                                                                                       Heeschen et al. (1986);
                                                                                                       Lorenz & Neumeier (1983)
                                    -              2709                  1.77                          Fooken & Butte (1987)

    Netherlands                     1983           278                   0.72 (0.27-2.20)b             Greve et al. (1985);
    (11 centres country-wide)                                                                          Greve & Wegman (1984)
                                    1977-1979,     2649                  2.1                           Olling (1984)

    United Kingdom (Scotland)       1979-1980      30                    0.01 (nd-0.04)                HMSO (1986)
                                    1983-1984      30                    < 0.01 (nd-0.02)

    Table 22.  (cont'd).

    Region                          Year           Number of             Mean concentration in         Reference
    Country                                        samples               mg/kg on fat basis (range)

    Italy (Rome)                    1983-1985      65                    0.070 (0.007-0.176)a,c        Dommarco et al. (1987)

    Finland (different parts)       1984-1985      183 (165 of           0.57 (0.05-10.7)              Mussalo-Rauhamaa et al.
                                                   women)                                              (1988)

    Sweden (5 regions)              -              300e                  1.06-1.18 (four regions)      Noren (1983)
                                                                                                       1.44 (one region)
                                    1972           227d                  1.05                          Noren (1988)
                                    1976           245                   0.99
                                    1980           340                   0.78
                                    1984-1985      102                   0.60

    Austria (Vienna)                -              22                    1.54 (0.58-3.78)              Pesendorfer (1975)

    Other regions                                  9                     1.29 (0.95-1.57)              Pesendorfer (1975)

    a  Whole milk.
    b  Median concentration.
    c  Arithmetic mean.
    d  Number of mothers that provided 4-7 samples each (samples were pooled).
    e  In each region, 300 mothers gave breast milk 3-5 days after parturition.

    In a study by Kuwabara et al. (1978), the relationship was
    investigated between breast-feeding and PCB residues in the blood of
    children whose mothers were occupationally exposed to PCBs. The
    children ingested their mother's milk for periods of < 1 to 3 years.
    The age of the children at the time of the study ranged up to 13
    years. The data provide evidence that PCBs are retained in the
    children's body for many years and that longer intake of mother's milk
    tends to increase PCB levels in the blood of the children. The PCB
    levels in the blood of the 20 occupationally-exposed women and their
    39 children ranged from 8.3 to 84.5 and 0.8 to 93.2 g/litre,

    The results suggest that the PCB levels in the blood of children are
    much more influenced by the transportation of PCBs through the
    mother's milk than through the placenta. Furthermore, it was found
    that the gas chromatographic patterns of the blood PCBs of the
    children, breast fed for a long time, were different from those of
    their mothers. Blood from 16 non-occupationally exposed mothers and
    their children (17), showed that, as the length of the breast-feeding
    period increased, there was an increase in the PCB levels in the blood
    of the children. The mean blood PCB level in mothers was 2.8 
    0.8 g/litre; in children, it was 3.8  3.6 g/litre. In this study,
    no clear change in blood PCBs patterns between mothers and children
    was observed (Kuwabara et al., 1979).

    Samples of maternal blood, milk, and umbilical cord blood were
    collected from 43 mothers giving birth to their first or second child;
    all the mothers had lived in Oslo during the previous 2 years. Blood
    samples were collected immediately after delivery, either by Caesarean
    section (16 Norwegians) or normally (20 Norwegians and 7 immigrants).
    Subcutaneous fat samples were obtained during the operation. Samples
    of colostrum and milk were obtained 3 and 5 days postpartum. PCBs were
    found in 135 of the total 168 samples. In the Norwegian women and
    infants, PCBs were the major contaminants, whereas only traces of PCBs
    were found in the samples of immigrants. The average concentrations in
    the maternal serum, cord serum, colostrum, and breast milk of
    Norwegian women (Caesarean and normally delivered taken together)
    were: 10, 3-5, 18-21, 20-23 g/kg wet weight (Skaare et al., 1988).  Major PCB congeners in human milk

    Commercial PCB preparations consist of complex mixtures of
    environmentally stable compounds with a wide range of chlorine
    contents. PCBs are transferred to breast-fed infants with the fat of
    the mother's milk. Thus, infants nurtured on maternal milk are exposed
    to relatively high concentrations of the higher chlorinated PCBs in
    the short period preceding the full functioning of certain organs,
    e.g., the liver (Jensen, 1983b; Slorach & Vaz, 1983; Gezondheidsraad,

    Three major congeners were present in breast milk, e.g., PCB congener
    numbers 138, 153, and 180 (DFG, 1988).

    Slorach & Vaz (1983) reported that the GC patterns of PCBs in breast
    milk samples from different countries were similar. The peaks denoted
    146, 174, and 180 were dominant in the gas chromatograms. The total
    levels of PCBs and the concentrations of certain congeners in Swedish
    human milk, sampled in 1972-89, were studied by Noren et al. (1990).
    Minor changes in the distribution of the congeners were found over the
    period of study. The most abundant of the non- ortho coplanar PCBs in
    Swedish human milk was 3,4,5,3',4'-pentachlorobiphenyl (126), with
    levels decreasing from 0.35 g/kg milk fat (1972) to about 0.10 g/kg

    Safe et al. (1985a) analysed a sample of breast milk using the
    congener-specific PCB method and found the following major components:
    2,4,4'-trichloro-; 2,4,5,4'-tetrachloro-; 2,4,5,2',4'-pentachloro-;
    2,4,5,3',4'-pentachloro-; 2,3,4,5,2',5'-hexachloro-;2,4,5,2',4',5'-
    hexachloro; 2,3,4,5,2',3',4'-heptachloro-; and 2,3,4,5,2',4',5'-

    The major PCB congeners in the breast milk of Japanese women from the
    general population were: 2,4,4'-trichloro-; 2,4,3',4'-tetrachloro-;
    2,4,5,3',4'-pentachloro-; 2,3,4,2',3',4'-hexachloro-;2,3,4,5,2',4'-
    hexachloro-; and 2,3,4,5,2',4',5'-heptachlorobiphenyls. The congeners
    were present in 5% or more samples; a few other congeners were present
    in only 1-3% (Gyorkos et al., 1985; Jensen, 1983b).

    Sixty-eight breast milk samples collected in the Netherlands were used
    to determine the congener distribution. The indicator congeners,
    present in the highest concentrations, were: 2,4,4'-trichloro-,
    2,4,5,2',5'-pentachloro-, 2,4,5,3',4'-pentachloro-,
    2,3,4,2',4',5'-hexachloro-, 2,4,5,2',4',5'-hexachloro-,
    2,3,4,5,2',4',5'-heptachlorobiphenyl (Wegman & Berkhoff, 1986).

    Schecter et al. (1989a) analysed a total of 17 samples of human milk
    from Thailand and Vietnam, for the presence of PCB congeners. The main
    congeners that were present were 138, 153, and 180 (each in the range
    of 8-31 g/litre). The other congeners, normally present, were all
    below the detection limit of 2 g/litre.

    In a study on pooled human milk samples from a 1982 nation-wide survey
    in Canada, Mes & Marchand (1987) compared the relative amounts of 29
    selected PCB isomers with amounts in milk samples of unexposed Rhesus
    monkeys. In the pooled milk sample, 397 g PCBs/litre, on a fat basis,
    were found and the PCB isomer numbers 74, 99, 118, 138, 153, and 180
    were the main contributors. Most of the predominant PCB isomers in
    human milk were also observed in monkey's milk, but monkey's milk had
    relatively low levels of PCB isomers numbers 74 and 99.

    In another study, Davies & Mes (1987) analysed breast milk samples
    from Canadian, Indian, and Inuit (Eskimo) mothers in Canada. The 18
    samples were received from 5 Indian and Inuit nursing zones. The
    combined total PCB isomer level (on a whole-milk basis) of the native
    population was comparable with that of the national population. Even
    the levels of the 5 largest PCB congeners (Nos. 74, 118, 138, 153, and
    180) were comparable.

    Individual congeners in the blood of Yusho- and Yu-Cheng patients are
    discussed in section 5.6.  Factors that influence the intake of PCBs with milk

    Present data suggest that the PCB content of human milk varies
    considerably from individual to individual.

    Many factors affect the level of PCBs and other organochlorine
    compounds in breast milk including the fat content of the milk; time
    from start of lactation; mother's age; mother's body weight; parity;
    number of children previously breast-fed; origin and residence; eating
    habits; season; smoking; use of household products; amount of milk;
    and exposure at work (WHO/EURO, 1985, 1988).

    In a given woman's milk, there are fluctuation in the PCB levels in
    whole milk and in milk fat during one nursing session and during the
    day (Jensen, 1983b). A decrease of PCB levels in both milk and milk
    fat has been found during the lactation period. Furthermore, the PCB
    concentration in human whole milk and milk fat increases with the age
    of donor. Another confounding factor is that the PCB levels decrease
    with increasing numbers of deliveries and lactations (Greve et al.
    1985); lactation serves as a period for the biological elimination of
    PCBs (Jensen, 1983b). The PCB levels in human milk are higher in
    heavily populated and industrialized areas than in rural areas.
    Furthermore, in general, the PCB levels in the breast milk of women
    from developing countries are lower (Jensen, 1983b).

    Cetinkaya et al. (1984) studied the PCB levels in human milk samples
    from all over the Federal Republic of Germany. At the same time, data
    were collected by means of a detailed questionnaire on residency,
    workplace, smoking, drinking and eating habits, and the age of
    participating individuals.

    The breast milk of 45 women consuming lacto-vegetarian food was
    compared with that of 41 women consuming conventional food in the
    Federal Republic of Germany in the period 1979-81. The PCB
    concentration was comparable, e.g., 2.2 and 2.5 mg/kg, on a fat basis,
    respectively (Acker et al., 1984).

    Fish consumption was positively correlated with PCB levels in maternal
    serum and breast milk. PCB levels in serum increased with age, but
    were unrelated to social class, parity, or body weight (Schwartz et
    al., 1983).

    Eight hundred and one Wisconsin anglers were surveyed for fishing and
    consumption habits in 1985. The mean annual number of sport-caught
    fish meals was 18 (range 7.1 to 33.3). The mean number of
    non-sport-caught fish meals was 24. The median PCB serum congener sum
    level for 192 anglers was 1.3 g/litre (range, nd to 27.1 g/litre).
    Statistically significant positive Spearman correlations were observed
    between sport-caught fish meals and PCB levels in serum and between kg
    of fish caught and PCB levels in serum (Fiore et al., 1989).

    PCBs were measured in maternal serum, cord blood, placenta, and serial
    samples of breast milk and colostrum, from 868 women in North Carolina
    (USA). Forty-three per cent of the women were primiparous. Breast milk
    was collected at 6 weeks, 3 months, and 6 months, and, in a few cases,
    up to 18 months postpartum. The median PCB concentration in breast
    milk decreased during the sampling period from 1.77 to 1.02 mg/kg, on
    a fat basis. The PCB concentration dropped by about 20% over 6 months
    and 40% over 18 months. This implies that excretion in milk is a major
    factor in lessening the mother's body burden; however, it also implies
    substantial exposure of the child. Colostrum contained a median value
    of 1.74 mg/kg. PCBs concentrations were higher in milk than in serum
    and higher in maternal serum than in the placenta. The levels in cord
    blood were almost always below the limit of quantification. Older
    women and women who regularly drank alcohol had higher PCB levels in
    their milk; blacks had higher levels than whites. In general, women
    had higher levels in their first lactation and in the earlier samples
    of a given lactation, and levels declined both with time spent
    breast-feeding and with number of children nursed (Rogan et al.,

    Two hundred and forty-two newborn infants of mothers who consumed
    moderate quantities of contaminated lake fish and 71 infants whose
    mothers did not eat such fish were examined during the immediate post
    partum period. PCB exposure was correlated with lower birth weight and
    smaller head circumference, and the authors claimed that these effects
    were not attributable to any of 37 potential confounding variables,
    including socioeconomic status, maternal age, smoking, etc. (Fein et
    al., 1984).

    The mother's diet may be an important determinant of the PCB levels in
    her milk. In some areas of the world, the intake of PCBs from eating
    contaminated fish has been claimed to be the most important source of
    PCBs in human milk. Dairy products and meat may be contaminated via
    natural food or feedstuffs (WHO/EURO, 1988).

    In a pilot study on the course of the PCB concentration in human milk
    during 6 months of lactation, some PCB determinants were studied in 23
    women and their infants. The average PCB concentration in the milk of
    14 mothers during a 6-month period amounted to 0.66  0.12 mg/kg, on a
    fat basis. In univariate analyses, the PCB concentration on a fat
    basis was strongly associated with pre- versus post-pregnancy weight
    gain, age, and occupation. After multiple regression analysis, the PCB
    concentration on a fat basis remained significantly associated with
    changes in weight gain. The pre-pregnancy Quetelet Index of the mother
    (height/weight) and the estimated PCB content of the diet (fish) were
    correlated with the PCB concentration, on a milk basis (Drijver et
    al., 1988).

    5.5.4  Other tissues

    Schecter et al. (1989b) analysed the tissues of 3 patients from the
    North American continent, with no known history of chemical exposure,
    for the presence of PCB isomers. The total PCB concentrations in the 9
    tissues studied were different. The highest levels were found in
    adipose tissue, subcutaneous fat (range 86-423 g/kg), adrenals
    (25-103 g/kg), liver (3-149 g/kg), bone marrow (26 g/kg), kidneys
    (2-31 g/kg); levels in the spleen, lung, and testes were below
    12 g/kg. Congeners present in the highest concentrations were numbers
    28, 74, 118, 153, 105, 138, 183, and 180.

    5.6   Accidental exposures (Yusho- and Yu-Cheng)

    In 1968, a large number of persons in Japan were accidentally poisoned
    by the consumption of a batch of rice oil contaminated with Kanechlor
    400. A similar accident happened in the Province of Taiwan in 1979,
    where the affected persons had also consumed rice-bran oil
    contaminated with PCBs. The 2 cases of poisoning were called Yusho and
    Yu-Cheng accidents, respectively (see section

    The average PCB concentration in the plasma of Yusho children was
    6 g/litre, compared with 3.7 g/litre in controls. Breast-fed Yusho
    children had higher levels than children not breast-fed (Abe et al.,

    The concentrations of PCBs in the adipose tissue, liver, and blood of
    Yusho patients, about 5 years after the outbreak, were 1.9 
    1.4 mg/kg, 0.08  0.06 mg/kg, and 6.7  .3 g/litre, respectively.
    These values were only about twice those of controls. The mean blood
    PCB level of 278 persons involved in the Yu-Cheng accident was
    89.1 g/litre (range 3-1156 g/litre). Six months after the exposure,

    the concentrations of PCBs in the blood had decreased to
    12-50 g/litre. The mean blood concentration of 165 patients,
    9-18 months after the onset of poisoning, was 38 g/litre (range
    10-720 g/litre) (see section The blood PCB level of some
    Yu-Cheng patients (99  163 g/litre), was much higher than that of
    the Taiwanese population (1.2  0.7 g/litre), one year after the
    outbreak of the intoxication.

    Chen et al. (1985) analysed the blood of 165 Yu-Cheng patients, 9-18
    months after the onset of poisoning, and found 10-720 g PCBs/litre
    with a mean value of 38 g/litre. The blood of 10 patients, 9-27
    months after poisoning, contained 0.02-0.2 g PCDFs/litre. The
    PCDF-congeners found in tissues were the same as those found by Masuda
    et al. (1985).

    Seven PCB congeners including: 2,4,5,3',4'-pentachloro-; 2,3,4,3',4'-
    pentachloro-; 2,4,5,2',4',5'-hexachloro-; 2,3,4,2',4',5'-hexachloro-;
    2,3,4,5,3',4'-hexachloro-; 2,3,4,5,2',4',5'-heptachloro-; and
    2,3,4,5,2',3',4'-heptachlorobiphenyls, were identified in the blood
    and tissues of Yusho, Yu-Cheng patients and controls.

    Major PCDF congeners identified in the tissues and blood of Yusho and
    Yu-Cheng patients were 2,3,6,8-tetrachloro-; 2,3,7,8-tetrachloro-;
    1,2,4,7,8-pentachloro-; 2,3,4,7,8-pentachloro-; and 1,2,3,4,7,8-
    hexachlorodibenzofurans. The 2,3,4,7,8-pentachloro-compound was
    predominant. The concentrations of PCDFs in the adipose tissue and
    liver of Yusho patients were 6-13 g and 3-25 g/kg tissue,
    respectively. No PCDFs could be detected in the controls. Besides PCBs
    and PCDFs, 4-methylthio-2,5,2',5'-tetrachlorobiphenyl (concentrations
    ranging from 0.1 to 1.4 g/kg tissue) and 4-methylsulfone-
    2,5,2',5'-tetrachlorobiphenyl (range 0.3-2.5 g/kg tissue) were also
    found (Masuda et al., 1985).

    5.7  Occupational exposure

    5.7.1  Accidental exposure

    Though the volatility of the PCBs is low, they are found in rather
    high concentrations in the workroom air in both the long-term open use
    of PCBs and in temporary or acute events where evaporation into the
    air is possible. The measured air concentrations of PCBs in long-term
    exposure situations, such as the manufacturing of transformers or
    capacitors, varied from 30 to 1000 g/m3, depending on the year of
    measurement and the factory concerned (Silbergeld, 1983).

    In discontinuous work, such as the inspection and repair of
    transformers and capacitors, levels of between 0.1 and 60 g/m3 have
    been observed (Wolff, 1985). PCB concentrations in the breathing zone
    of workers in transformer repair and maintenance work varied between
    0.01 and 24.0 g/m3 (Moseley et al., 1982).

    In the atmosphere of an electroindustrial plant in Bela Krajina,
    levels in the manufacturing room, where the autoclave was emptied,
    averaged 2000 g/m3 (range, 1400-3200  g/m3); an average of
    80 g/m3 (range 40-120 g/m3) was found in the working environment
    in capacitor manufacture (Jan et al., 1988b).

    Digernes & Astrup (1982) determined the concentrations of PCBs in the
    atmosphere of the workplace of data screen operators, because skin
    rashes and eczema had been reported among the workers. The PCB
    concentrations in the working atmosphere (3 samples: concentrations
    ranging from 0.056 to 0.081 g/m3) were about 50-80 times higher than
    the maximum level of PCBs in 3 samples collected outside the building
    (0.0005-0.001 g/m3). The indoor and outdoor samples also differed
    qualitatively. The indoor samples contained only Aroclor 1242, while
    outdoor samples contained a mixture of Aroclor 1242 and 1254.

    Acute emergency events may cause extremely high concentrations of PCBs
    in the air, particularly in cases when PCBs are burnt or heated (fire,
    short circuit with electric arcing, burning in welding, etc.). Levels
    of up to 10 000-16 000 g/m3 have been measured. In the case of
    extensive leaks of unheated PCBs from capacitors, concentrations of
    1900 g/m3 have been measured in workroom air (Elo et al., 1985;
    WHO/EURO, 1987).

    In connection with fires and electrical explosions, due to short
    circuits, PCBs may be decomposed at elevated temperatures varying from
    a few hundred to 2000C. Soot may be produced in large amounts,
    consisting of particles that may contain PCB concentrations up to
    5000-8000 mg/kg of soot (Elo et al., 1985; O'Keefe et al., 1985;
    WHO/EURO, 1987).

    When evaluating PCB exposure, it is important to take into account
    skin absorption from surfaces and tools, in addition to exposure via
    inhalation. Surface concentrations of PCBs in capacitor factories have
    varied between 4 and 60 g/m2, and, where PCB leaks have occurred,
    levels of up to 30 mg/m2 have been measured. Where PCBs have been
    used long-term, contamination levels of 1-2 g/cm2 have been found on
    tools and tables.

    A transformer was found to have overheated and released an oily mist
    containing PCBs and their pyrolysis by-products, in a Department
    building in New Mexico. The transformer contained Askarel (87% Aroclor
    1260 and 13% of a mixture of tri- and tetrachlorinated benzenes). The
    3-storey building was extensively contaminated via the following ways:

    *   mist entered 2 rooms, adjacent to the basement in which the
        transformer was located;

    *   direct spread of mist and fumes through stairways;

    *   air drafts created by open windows and exhaust fans, spreading
        fumes throughout the building;

    *   foot traffic by employees and other persons;

    *   the exhaust vent of the transformer room, located near the intake
        vents for the building's air-conditioning system.

    Air samples obtained up to 14 h after the incident showed levels of
    48 g/m in the transformer vault and 20 g/m3 in the room above the
    vault. Wipe samples of surfaces showed PCB levels ranging from 30
    million g/m2 for grossly contaminated surfaces to 4700 g/m2 for
    surfaces without visible contamination.

    Five to 7 days later, air and surface samples were analysed for
    2,3,7,8-tetrachlorodibenzofuran (TCDF), which was found to be present
    in the air at an average level of 48 g/m3 in most contaminated
    areas. In wipe samples, the levels ranged from 5 ng/m2 to
    41.224 ng/m2. 2,3,7,8-Tetrachlorodibenzo- p-dioxin (TCDD) was not
    detectable in either air samples (detection limit, 0.5-5.0 pg/m3 air)
    or wipe samples (detection limit 180 ng/m2) (Anon., 1985).

    Very high concentrations of these toxic chemicals may be found in soot
    emitted in connection with fires and explosions in capacitors.

    Thus, skin contamination, and the ingestion and inhalation of soot
    particles, may result in serious exposure in PCB accidents and

    A short-term, follow-up study was performed on 55 workers in a gear
    plant, whose work did not involve the use of PCBs. Exposure was to the
    total residual PCB left behind by a capacitor company that had
    formerly (3 years before) used the site. Air samples contained
    < 10 g/m3 and mean concentrations in wipe samples ranged from 23 to
    161 g/100 cm2. The 38 workers had a mean PCB concentration in serum
    of 14.4 and the 17 office workers, 4.8 g/litre. When the PCB
    determinations were repeated in the 2 following years, no clear
    decrease was observed (Christiani et al., 1986).

    5.7.2  Occupational exposure during manufacture and use

    Occupational exposure occurs during the manufacture of PCBs as well as
    during their use by the electrical industry. It may also be widespread
    among mechanics in contact with lubricating oils and hydraulic fluids,
    among workers exposed to varnishes and paints, and among office
    workers who have contact with pressure-sensitive duplicating paper
    (carbonless copying paper), some brands of which readily transferred
    PCBs to skin (Kuratsune & Masuda, 1972).  Adipose tissue

    Levels of PCBs in the adipose tissue of occupationally exposed workers
    have been found to vary between 26 and 50 mg/kg (range, 2.2-290 mg/kg).
    There is a strong correlation between the blood PCB concentration and
    PCB levels in adipose tissue, but the distribution of the various
    congeners between plasma and adipose tissue is not the same, as
    described above.

    Emmett (1985) found the following congeners in the adipose tissue of
    present and past transformer workers exposed to Aroclor 1242 and 1254:
    2,4,3',4',5'-pentachloro-, 2,3,4,3',4'-pentachloro-, 2,3,4,5,2',4'-
    hexachloro-, 2,3,4,6,3',4'-hexachloro-, 2,4,5,3',4',5'-hexachloro-,
    2,3,4,5,2',3',4'-heptachloro-, and 2,3,4,5,6,3',4'-hepta-
    chlorobiphenyl.  Blood

    Karppanen & Kolho (1973) analysed the blood of 26 persons, 9
    non-exposed, 6 persons handling PCBs, and 11 persons employed for 4
    years in a capacitor-manufacturing plant in Finland. In the latter
    case, Aroclor 1242 was used. The average concentrations in the blood
    of the 3 groups were 7.1 g/kg (3.1-12 g/kg), 49.5 g/kg
    (36-63 g/kg), and 440 g/kg (70-1900 g/kg), on a wet weight basis.

    More recent results of a Finnish control group of workers indicated
    serum PCB levels of 1.2  0.6 g/litre in an industrial area (Luotamo
    et al., 1985; WHO/EURO, 1987). With acute exposure to high
    concentrations of PCBs in air (8000-16 000 g/m3), for a short
    period, blood PCB concentrations rose to levels of 30 g/litre; a
    return to the normal level of 3 g/litre was achieved, 4 weeks after
    termination of exposure (Elo et al., 1985; WHO/EURO, 1987).

    Similar plasma values were found in workers from Japanese capacitor
    factories, but, here, skin lesions were noted (Hasegawa et al.,
    1972a). In this same study, it was reported that air levels of PCBs of
    10-50 g/m3 were measured in a factory where KC-300 was used in the
    manufacture of electric condensers. PCB levels in the serum of workers
    ranged from 100 to 650 g/litre. One month after the use of PCBs had
    been suspended, serum levels remained unchanged (90-740 g/litre).
    However, in another factory making electric condensers, serum levels
    decreased from an average of 800 to 300 g/litre, within 3 months of
    the use of PCBs being discontinued (Kitamura et al., 1973). According
    to Hara et al. (1974), the half-time of PCBs in the blood of workers,
    engaged in the manufacture of electric condensers for less than 5
    years, was several months, while that of workers employed for more
    than 10 years was 2-3 years.

    Kuwabara et al. (1978) reported mean PCB levels of 36.8 g/litre
    (range 8.3-84.5 g/litre) blood in 20 PCB-workers, 39 children had
    blood levels of 14.3 g/litre (0.8-93.2 g/litre), and 12 Yusho
    patients, 4.2 g/litre (1.8-8.6 g/litre).

    Fact-finding surveys of 63 workers, who were occupationally exposed to
    PCBs (Kanechlor 500) in the production of silk thread or of paint,
    were carried out in Japan in 1974-75; some of them and their families
    were also surveyed again in 1975-82. Nineteen per cent of them showed
    PCB levels higher than 50 g/litre plasma. These persons did not show
    the typical clinical findings of Yusho patients. During 7 years, no
    clear decline was observed (Takamatsu et al., 1984).

    There is clear evidence that relatively high PCB levels persist in the
    blood of workers whose "external" exposure ceased several months or
    years previously. The blood PCB concentrations in capacitor
    manufacturing workers, who had been exposed for 1-24 years, varied
    between 24.4 and 192 g/litre; this was higher than levels in the
    blood of a reference population (0.5-33 g/litre) (Maroni et al.,

    In Japan, Yakushiji et al. (1984a) studied the rate of decrease and
    the half-life of PCBs in the blood of children (aged 1-13 years) and
    their mothers, who were occupationally exposed to PCBs, over a 5-year
    period  (1975-79). The mean concentration of 121 blood samples from
    50 children was 17.4  22.9 g/litre and that in 65 samples from 29
    mothers was 32.3  20.6 g/litre. The concentrations of PCBs in the
    blood of the children varied over a wide range, because of differences
    in the duration of breast-feeding. The rate of decrease of the PCB
    concentration in the blood in both 18 children and 8 mothers was
    relatively constant and independent of the PCB concentrations. A
    one-compartment model equation was sufficient to represent the
    decrease in the concentration of PCBs in the blood. The mean rate
    constant of the decrease for the children was 24.2% per year,
    approximately 2.6 times higher than that of the mothers (9.2%),
    equivalent to half-lives of 2.8  1.1 and 7.1  2.7 years,
    respectively. The dilution effect due to the increase in body weight
    was the most important factor that affected the reduction of the PCB
    concentrations in the children.

    A total of 118 blood samples, mainly from employees in industries
    using PCBs, were collected in the period 1975-85. In 64 blood samples,
    an average level of 17 g/litre (range nd-110 g/litre) was found
    (Frank et al., 1988).

    Brown & Lawton (1984) studied the partitioning of PCBs between adipose
    tissue and serum in a population of 173 capacitor workers, who were
    occupationally exposed to Aroclors 1254, 1242, and 1016 for various
    periods of time. The serum levels of PCBs were significantly dependent
    on the level of lipids in the serum, but not on that in the albumin.
    The apparent contribution of cholesterol and its esters to PCB
    transport is nearly equal to their contribution to the total serum
    neutral lipids. The level of serum lipids PCBs must be equal to the
    adipose fat PCBs level.

    Yakushiji et al. (1984b) studied the relationship between
    breast-feeding and the PCB levels in the blood. The blood samples of
    50 children (121 samples) and of 29 occupationally exposed mothers (65
    samples) were analysed during the period 1975-79. The PCB levels in
    the blood of the children were greatly influenced by the duration of
    breast-feeding, but showed little relationship to the PCBs levels in
    maternal blood.


    6.1  Absorption

    6.1.1  Inhalation

    Studies on rats (6 per group) showed that an aerosol containing a PCB
    mixture (Pydraul A200: 42% chlorine), particle size 0.5-3.0 m, at a
    concentration of 30.4  3.4 g/m3 for 30 min, was readily absorbed
    through the lungs. The PCB concentration in the liver, 15 min after
    cessation of exposure, was 50% of the maximum concentration attained
    after 2 h (70 mg/kg tissue) (Benthe et al., 1972).

    6.1.2  Dermal

    Vos & Beems (1971) and Vos & Notenboom-Ram (1972) applied Aroclor 1260
    to the shaved backs of rabbits and found systemic effects in the
    kidneys, indicating that PCBs can penetrate the skin (see section

    Nishizumi (1976), using tritium-labelled PCBs (40% chlorine), found
    evidence for the dermal absorption of PCBs in rats.

    In a study of the occupational exposure of electrical workers to PCBs
    (Pyralen 3010 and Apirolio, 42% chlorine content), Maroni et al.
    (1981a) concluded that absorption of PCBs occurred through the human
    skin. Quantitative data were not available.

    6.1.3  Oral

    When polychlorobiphenyl isomers were administered orally, by gavage,
    to rats, at levels of 5, 50, or 100 mg/kg body weight for the lower
    chlorinated compounds and up to 5 mg/kg for the higher chlorinated
    compounds, 90% of the compounds were rapidly absorbed by the
    gastrointestinal tract (Albro & Fishbein, 1972; Berlin et al., 1973;
    Melvs & Brandt, 1973).

    Using Rhesus monkeys, Allen et al. (1974a,b) determined that > 90% of
    a single oral dose of 1.5 or 3.0 g Aroclor 1248/kg body weight was
    absorbed over a period of 2 weeks. Drill et al. (1981) and US EPA
    (1985) reviewed a number of studies indicating that PCBs are readily
    absorbed from the gastrointestinal tract following oral

    Bleavins et al. (1984) found that, over a period of 5 weeks, European
    ferrets absorbed 85.4% of a single dose of 14C-labelled Aroclor 1254
    (0.05 mg) given in food.

    In contrast to the above studies, Norback et al. (1978) claimed that
    59.3-87% of a single oral dose of 2,4,5,2',4',5'-hexachlorobiphenyl
    passed unabsorbed through the intestines of monkeys, the first week
    after dosing.

    6.2  Distribution

    6.2.1  Inhalation (rat)

    Maximum PCB concentrations in the liver and brain of rats occurred 2
    and 24 h, respectively, after a single, 30-min exposure to 30.4 
    3.4 g/m3 of Pydraul A200 aerosol (42% chlorine content). The
    concentrations in these tissues declined, while concentrations in
    adipose tissues reached a maximum after 48 h (Benthe et al., 1972).

    6.2.2  Oral (rat)

    As in the case of other lipophilic substances, the absorption and
    distribution of PCBs will, in all probability, take place via the
    lymphatic system (by the chylomicrones) (DFG, 1988).

    Following absorption, the clearance of PCBs from the blood and tissues
    follows a biphasic pattern. The compounds rapidly clear from the blood
    and accumulate in the liver and adipose tissue or are metabolized in
    the liver to metabolites that are excreted in the urine and/or bile
    (Drill et al., 1981).

    Kurachi & Mio (1983) exposed mice to Kaneclor 400 at 100 mg/kg diet,
    for 5-20 days. High levels were found in the gonads, skin, adipose
    tissue, adrenals, and kidneys.

    In a study by Grant et al. (1971a), 4 days after an oral dose of
    Aroclor 1254 was given to rats at 500 mg/kg, the concentrations of
    PCBs in the fat, liver, and brain were 996, 116, and 40 mg/kg,
    respectively. Similar results showing that the highest concentration
    was in the fat, were obtained in rats given Aroclor 1254 in the diet
    (Curley et al., 1971), in boars (Platonow et al., 1972), cows
    (Platonow & Chen, 1973), and in pigeons and quail (Bailey & Bunyan,
    1972). In the studies of Curley et al. (1971), the tissue
    concentrations initially showed a rapid rise and then a slow increase
    while the PCB diet was being administered; Grant et al. (1974) fed
    diets containing Aroclor 1254 at 0.2, 20, and 100 mg/kg to rats for 8
    months, during which period the tissue concentrations reached a steady
    state that was dose-dependent (Table 23). Similar tissue distribution
    data for Aroclors 1016 and 1242 have been reported by Burse et al.
    (1974) and for Kanechlor-400 by Yoshimura et al. (1971).

        Table 23.  Tissue distribution of PCBs (mg/kg wet weight) in rats fed Aroclor 1254,
               Aroclor 1242, or Aroclor 1016 at 100 mg/kg for about 6 months

    Tissue          Aroclor 1254a         Aroclor 1242b         Aroclor 1016b

    Blood           0.40                  0.53 (plasma)         0.38 (plasma)
    Liver           16                    4.21                  7.86
    Brain           3.4                   1.69                  2.98
    Kidneys                               1.89                  3.21
    Heart           7.3
    Fat             32.0                  110                   236
    Urine                                 0.03                  0.28

    a  From: Grant et al. (1974).
    b  From: Burse et al. (1974).

    The study by Burse et al. (1974) showed that, with continuous feeding
    of 3 types of Aroclor (see Table 23) at 100 mg/kg diet, a steady state
    was not reached for 6-8 months and that the decline of stored PCBs in
    adipose tissue (when the animals were kept on a PCB-free diet) was
    slow and did not reach zero during a recovery period of 5-6 months.
    This is surprising because the Aroclor sample used should not have
    contained appreciable amounts of hexachlorobiphenyls and higher
    isomers. Mizutani et al. (1977), discussing this aspect, came to the
    conclusion that mobilization from storage sites rather than metabolism
    constitutes the rate-limiting step in the depletion of the body burden
    of PCBs.

    It was demonstrated that, as the number of chlorine atoms on the
    biphenyl rings increased from 1 to 6, the tissue/blood ratio tended to
    increase. This increase was also proportional to the amount of lipid
    in the tissues with, consequently, a higher degree of bioaccumulation
    (Matthews, 1983, cf. WHO/EURO, 1988).

    The fat-plasma partition-coefficients for the different PCB congeners
    range from 50 up to 310 (DFG, 1988).

    6.2.3  Oral (monkey)

    Feeding studies were carried out on female rhesus monkeys given doses
    of 0, 5, 20, 40, or 80 g Aroclor 1254/kg body weight per day, for a
    period of 37 months (Arnold et al., 1984). Eighty monkeys were divided
    into 5 groups, each of 16 animals. The mean body weight of the monkeys
    at the start of the study was 6.44 kg. The Aroclor 1254 was dissolved

    in corn oil with glycerol as sweetener and fed to the monkeys in
    gelatin capsules. Samples of blood, adipose tissue, and faeces were
    collected every month and the presence of PCBs, determined. After 27
    weeks, levels of 1, 2-3, 9, 18, and 37 mg PCBs/kg fat were found in
    the 5 groups; after 47 weeks, blood levels were 1-3, 12, 35, 73, and
    129 g/litre respectively. PCB concentrations in whole blood increased
    more rapidly during the first 10 months of the study than in the
    remaining 27 months, in all groups. Concentrations in adipose tissue
    (fat) increased continuously during the 37 months. The ratio profiles
    of PCB levels in blood/adipose tissue, remained relatively static
    between the second and twenty-seventh month of feeding. The data in
    terms of relative concentrations (concentration/dose) suggest that the
    bioaccumulation or retention of PCBs may be dose-dependent,
    particularly for adipose tissue. The data available from PCBs in
    faeces indicate a dose-dependent PCB absorption.

    6.2.4  Oral (humans)

    According to the study by Nishimura et al. (1976) cf. Katsunuma et al.
    (1985), the PCBs within a human fetus are not evenly distributed. The
    concentrations of PCBs were highest in the skin and lowest in the
    brain among the 5 major organs (cerebrum, heart, liver, kidneys, and
    skin). That the highest level was found in the skin might have been
    because of the high solubility of the compounds in adipose tissue. In
    other words, PCBs accumulate increasingly as the body fat of a fetus
    increases. The authors stated that the low residue levels in the brain
    were likely to be because PCBs have a poor affinity for the brain

    6.2.5  Individual congeners of PCBs

    More detailed information on the tissue distribution of PCBs and their
    metabolites has been obtained by the administration of pure
    14C-labelled compounds, using both whole-body autoradiography and
    scintillation counting of tissue samples. Berlin et al. (1975)
    demonstrated that, after a single oral dose of 14C-labelled
    2,5,2',4',5'-pentachlorobiphenyl, radioactivity rapidly entered the
    circulation of mice and was distributed in the tissues, particularly
    in the liver, kidneys, lungs, and adrenals. Subsequently, the
    radioactivity in the body fat increased, rising to a maximum within
    4-24 h. In most other tissues, the radioactivity decreased rapidly
    after dosing, but the authors noted a special affinity for the skin,
    the bronchiolar epithelium of the lungs, and certain glandular
    secreting tissues. Soon after administration of the dose,
    radioactivity appeared in the bile and was eliminated in the faeces.
    Similar results were obtained by Melvs & Brandt (1973) in mice
    treated with 2,4,2',4'-tetrachlorobiphenyl, which possessed a high

    affinity for the adrenal cortex, the corpora lutea, and glandular
    secreting tissue. In quail treated with 2,4,2',3'- and
    2,4,3',4'-tetrachlorobiphenyl, the radioactivity in the egg yolk was
    high, exceeding that in the fat. Gage & Holm (1976) determined
    concentrations in the abdominal fat of mice, 7 and 21 days after they
    were administered a single dose (13-165 g/mouse) of one of 14 PCB
    congeners, by gavage. Relatively low levels (< 10 ng/g per g dose)
    were found at 7 days for 4,4'-dichloro-; 3,2',4',6'-tetrachloro, and
    2,3,4,2',4',6'-hexachlorobiphenyl with relatively high levels (>
    100 ng/g per g dose) for 2,4,5,2',4',5'-hexachloro-, and the
    4,2',4',6'-, and 2,4,2',4'- tetrachlorobiphenyls.

    Muehleback & Bickel (1981) treated rats, by gavage, with a single dose
    of 14C-2,4,5,2',4',5'-hexachlorobiphenyl at 0.6 or 3.6 mg/kg body
    weight. The rats were examined 1 h, 24 h, 6 weeks, 20 weeks, or 40
    weeks after dosing. The highest levels of PCBs were found in the
    muscle, liver, adipose tissue, and skin, early in the study. By the
    end of the study, the highest PCB levels were found in the adipose
    tissue followed by the skin, muscle, and liver. During the 40-week
    study period, only 16% of the total dose was excreted.

    The pharmacokinetics of individual monochloro-, dichloro-,
    tetrachloro-, pentachloro-, and hexachlorobiphenyls were studied by
    Matthews & Anderson (1975a,b), Lutz et al. (1977), and Tuey & Matthews
    (1977). The mono- and dichlorobiphenyls were largely removed from
    adipose tissue within 4-7 days, the 3 higher chlorinated biphenyls
    were eliminated much more slowly. The half-life for the
    tetrachlorobiphenyl from adipose tissue was 15 days. Skin effects were
    more or less comparable.

    Beran et al. (1983) studied the distribution of 14C-labelled
    2,5,4'-tri-, 2,4,5,2',4',5'-hexa-, and 2,3,4,5,2',3',4',6'-
    octachlorobiphenyl in the haematopoietic tissues of squirrel monkeys
     (Saimiri scureus) and C67Bl mice using whole-body autoradiography.
    An accumulation of radioactivity was observed in the bone marrow of
    one monkey after iv injection (substances dissolved in DMSO) of the
    tri- or hexachlorobiphenyl. The same was found in 3 normal mice
    treated with the octachlorobiphenyl. A study using whole-body
    autoradiography and spleen-colony assay in supralethally irradiated
    mice, implanted with syngenic bone-marrow cells, indicated that the
    major part of the radioactivity was localized outside the bone-marrow
    haemic compartment, probably in the fat. Nevertheless, the trichloro-
    and octachlorobiphenyls were found to inhibit the  in vitro formation
    of granulocytic colonies from mouse progenitor cells. Very low uptake
    of labelled chlorobiphenyls was observed in the thymus, spleen, and
    lymph nodes.

    14C-labelled-2,4,2',4'-tetrachloro- and 3,4,3',4'-tetrachlorobiphenyl
    were each administered orally to male Sprague-Dawley rats in a single
    dose at 0.54 mg/kg and 0.51 mg/kg, body weight, respectively.
    Distribution and covalent binding were studied. The accumulation of
    2,4,2',4'-tetrachlorobiphenyl in adipose tissue was much higher than
    that of 3,4,3',4'-tetrachlorobiphenyl, though the level in the blood
    was consistently higher in the 3,4,3',4'-tetrachlorobiphenyl-treated
    rats. The radioactivity bound in covalent linkages with cellular
    macromolecules in several tissues was determined. The data indicated
    that covalent binding was higher in 3,4,3',4'-tetrachloro-
    biphenyl-treated rats than in those treated with 2,4,2',4'-
    tetrachloro-biphenyl, particularly in the liver and blood components.
    These results suggest that the 2 tetrachlorobiphenyl isomers have
    different pharmacokinetic properties in rats and that the association
    of covalent binding with 3,4,3',4'-tetrachlorobiphenyl induced
    toxicities might be important. The microsomal enzyme system is likely
    to play an important role in the  in vivo covalent binding of
    tetrachlorobiphenyls (Shimada & Sawabe, 1984).

    In pharmacokinetic studies, 11 groups of 3 male ICR mice/group were
    administered daily doses of 100 mg 2,5,2',5'-tetrachlorobiphenyl/kg
    body weight dissolved in corn-oil/acetone (9:1), by gavage, for 8
    consecutive days. Thirteen groups of 3 mice were administered (by
    gavage) 8 mg 3,4,3',4'-tetrachlorobiphenyl/kg in the same vehicle
    every other day for 10 doses. One group was sacrificed just before
    each of the last 3 doses, the other groups were sacrificed at
    intervals of 0.5-336 h after dosing. After dosing to an apparent
    steady-state, 2,5,2',5'-tetrachlorobiphenyl was found to have a tissue
    elimination half-life of between 39.5 and 70 h. The half-life of
    3,4,3',4'-tetrachlorobiphenyl was 26-62.5 h. The 3,4,3',4'-
    tetrachlorobiphenyl had a substantially greater partitioning from
    serum into adipose tissue, liver, and thymic tissues. Studies were
    undertaken to compare the toxic potency of these 2
    tetrachlorobiphenyls, when similar tissue concentrations of the 2
    isomers were achieved in target and storage tissues. The studies
    demonstrated that thymic atrophy occurs at lower doses and tissue
    concentrations of 3,4,3'4'-tetrachlorobiphenyl than those required to
    produce hepatotoxicity. These two organ toxicities were produced only
    by 3,4,3'4'-tetrachlorobiphenyl, despite the fact that equivalent or
    higher tissue concentrations of 2,5,2',5'-tetrachlorobiphenyl were
    achieved  in vivo, in all tissues. The conclusion was that the
     in vivo  difference in the toxic potency of these tetrachloro-
    biphenyl isomers does not result from the differences in their tissue
    disposition, elimination, and ultimate bioaccumulation (Clevenger et
    al., 1989).

    6.2.6  Appraisal

    Matthews & Dedrick (1984), in a review, concluded that the
    pharmacokinetics of PCBs are complicated by numerous factors, not
    least of which is the existence of 209 different chlorinated
    biphenyls. While all PCB congeners are highly lipophilic and most are
    readily absorbed and rapidly distributed to all tissues, PCBs are
    cleared from the tissues at very different rates, and the same
    congeners may be cleared at different rates by different species. With
    the exception of special situations in which PCBs may be passively
    eliminated in lipid sinks, e.g., milk or eggs, clearance is minimal
    prior to metabolism to more polar compounds. Rates of PCB metabolism
    vary greatly with species and with the degree and positions of
    chlorination. Mammals metabolize these compounds most rapidly, but,
    even among mammalian species, the rates of metabolism vary greatly. In
    all species studied, the more readily metabolized chlorinated
    biphenyls have adjacent unsubstituted carbon atoms in the 3-4
    positions. Congeners that do not have adjacent unsubstituted carbon
    atoms may be metabolized very slowly and therefore cleared very
    slowly. PCBs that are not readily cleared concentrate in adipose

    6.3  Placental transport

    6.3.1  Laboratory animals

    The results of a number of animal studies have demonstrated that PCBs
    and specific congeners can cross the placental barrier and accumulate
    in the tissues of fetuses (US EPA, 1987). In studies in which monkeys
    were exposed prior to, and during, gestation, signs of PCB
    intoxication were observed in nursing, but not in newborn offspring
    (Allen & Barsotti, 1976; Iatropoulos et al., 1978). Results such as
    these have led to the conclusion that transfer through nursing may
    account for higher exposure of the young than placental transfer.

    Groups of pregnant ddN mice were fed diets containing Kanechlor 500
    (mainly comprising pentachloro- and hexachlorobiphenyls) at 0.01
    (controls), 0.94, or 86 mg/kg diet from day 1 to 18 of pregnancy.
    Regardless of the dietary level of PCBs, whole-body levels in the
    fetuses were only 0.1-0.2% of the total maternal intake, indicating
    limited transplacental transfer (Masuda et al., 1978a). Two groups of
    ddN mice were fed Kanechlor 500 (mainly containing pentachloro- and
    hexachlorobiphenyls at 0 or 0.94 mg/kg diet) from the day of
    insemination throughout gestation and for 5 weeks after delivery of
    offspring. Total PCBs were 100 times greater in the suckling animal
    than in the fetuses at term, from dams fed the same amount, indicating
    a considerable transfer of PCBs during lactation (Masuda et al.,

    Masuda et al. (1979) fed female ddN-mice diets containing
    polychlorinated biphenyls: 2,4,4'-trichloro-; 2,5,3',4'-tetrachloro-;
    2,4,5,2',5'-pentachloro-; 2,3,4,2',4',5'-hexachloro-;2,4,5,2',4',5'-
    hexachloro-; 2,3,4,5,6,2',5'-heptachloro-; and 2,3,4,5,2',3',4',5'-
    octachlorobiphenyl at levels of 0.32, 0.42, 0.42, 0.44, 0.44,
    0.16, and 0.23 mg/kg diet, respectively, for 18 days prior
    to, or after, mating. Animals were either sacrificed on day 18 of
    gestation or allowed to deliver and the offspring maintained for 5
    weeks on a normal diet. All the PCBs were qualitatively transferred
    across the placenta and through the milk. The amount transferred
    during lactation was greater than that transferred transplacentally.

    The transfer of 2,4,5,2',4',5'-hexachlorobi[14C-]phenyl across the
    placenta during the course of pregnancy in Sprague-Dawley mice was
    studied by Vodicnik & Lech (1980). The PCB was injected
    intraperitoneally at 100 mg/kg body weight, in corn oil, 2 weeks prior
    to mating. The concentrations of 14C-PCB in the fetuses from 12- and
    18-day pregnant animals were 0.71 and 2.45 mg/kg tissue, respectively.
    At birth, the total carcass concentration for all newborn animals was
    less than 3 mg/kg tissue, which represents less than 3% of the dose
    present in the mothers at birth.

    Placental transfer of polychlorinated biphenyls has also been reported
    in the mouse by Berlin et al. (1975) and Melvs & Brandt (1973).

    Curley et al. (1973a) found some placental transport of Aroclor 1254
    in the rat.

    Groups of pregnant and non-pregnant Wistar rats received a dose of
    14C-2,4,5,2',4',5'-hexachlorobiphenyl (2.1 C/kg), intraperitoneally.
    The amount of radioactivity transferred through the placenta was 2.7%
    of the administered dose, whereas 39.2% of the original dose was
    transferred through the milk (Ando et al., 1978).

    Aroclors 1221 and 1254 were found to cross the placenta of rabbits,
    when administered orally to does during gestation. The concentration
    in fetal tissues was dose-dependent and much lower with Aroclor 1221
    than with Aroclor 1254; the concentration of the latter in the fetal
    liver was greater than that in the maternal liver (Grant et al.,

    Bleavins et al. (1984) fed female European ferrets a single dose of
    14C-labelled Aroclor 1254 in the diet (0.05 mg), early (day 14) or
    late (day 35) in gestation, and determined the placental transfer of
    PCBs. Placental transfer to the kits was 0.01% (per kit) of the
    maternal dose, when the dams were exposed early in gestation, and,
    0.04%, when the dams were exposed late in gestation. Placental
    transfer of PCBs was considerably less than mammary transfer, with a
    ratio at 1 week of lactation of 1:15 and 1:7 for offspring of dams
    dosed early or late in gestation, respectively.

    Groups of lactating mother Rhesus monkeys, between 1 and 3 months post
    partum, received 16 mg Clophen A-30/kg per day for 30 days. One
    mother/infant pair served as a control. Clophen A-30 concentrations in
    the serum of both mother and infant and the milk were determined on
    days -14, -7, 0, 1, 2, 4, 8 and at weekly intervals thereafter. One
    mother and all infants were killed and tissues taken for PCB analysis.
    The concentration of Clophen A-30 in milk was 20 times higher than
    maternal serum levels. Infant serum levels were 2-5 times higher than
    their mothers. Tissue levels were generally higher in the infants.
    Clophen A-30 tended to concentrate in the infant fat, bone marrow, and
    adrenals (Bailey et al., 1980).

    Groups of 24 Rhesus monkeys were maintained on diets that provided
    Aroclor 1016 at doses of 0, 4.5, or 18.1 mg/kg body weight per day
    throughout gestation and a 4-month nursing period. At birth, the
    concentrations of PCBs in the skin of infants were similar to
    concentrations in the subcutaneous fat of the mothers. At weaning, the
    PCB content in the mesenteric fat of the infants was 4-7 times greater
    than that in the subcutaneous fat of the mothers. Gas chromatographic
    patterns showed that the adult adipose tissue did not include the
    total spectrum of peaks observed in the Aroclor 1016 standard, and
    that all of the peaks in the mesenteric fat of the infants at weaning
    and 4 months after weaning were qualitatively similar to those in the
    adult adipose tissue. According to the authors, these data suggested
    an inability of the fetus to metabolize and excrete certain congeners
    that are more readily metabolized and eliminated by adults and older
    infants (Barsotti & Van Miller, 1984).

    6.3.2  Wildlife

    A 6 1/2-year-old desert bighorn  (Ovis canadensis cremnobates) ewe
    and her term ram fetus were used to study the distribution and
    concentrations of PCBs in different organs and tissues. Fourteen
    maternal and 13 fetal tissues were analysed for their presence of
    organochlorine hydrocarbons. PCBs averaged 85 and 88% of the total
    residue loads for maternal and fetal tissues, respectively. It is
    remarkable that the "natural" PCB levels in the different organs and
    tissues were nearly the same, i.e., in maternal organs and tissues
    between 0.37 and 0.44 mg/kg, and, in fetal organs and tissues, between
    0.30 and 0.35 mg/kg, on a fat basis (Turner, 1979).

    6.3.3  Humans

    Four studies of placental passage in humans, based on small samples
    drawn from the general Japanese population, have yielded inconsistent
    results (Yoshimura, 1974; Akiyama et al., 1975; Kodama & Ota, 1977;
    Masuda et al., 1978a).

    PCBs were detected in the umbilical tissues, umbilical blood, amniotic
    fluid, and baby's blood from a woman who was occupationally exposed to
    Kanechlors 300 and 500 in a capacitor factory (Yakushiji et al.,
    1978). PCB levels in these tissues and fluids were considerably lower
    than that in the mother's blood.

    Jacobson et al. (1984b) examined maternal and cord serum (196 or 198
    samples each) for the presence of PCBs, in women who resided in the
    Michigan area (USA), where, in 1973, a PBB-incident occurred. Mean
    concentrations of maternal and cord serum were 4.7 g/litre (1.1-
    14.3 g/litre) and 2.0 g/litre (0.1-7.2 g/litre), respectively.
    Placental passage was indicated by a significant maternal to cord
    serum correlation for PCBs. The fact that cord serum levels were lower
    than those in maternal serum is consistent with the notion that the
    placenta may function as a partial barrier. The transfer rate of PCBs
    in maternal blood through the placenta to cord blood may vary,
    depending on the chemical nature of each PCB isomer (Ando et al.,
    1984). Ando et al. (1985) examined the PCB concentrations in the
    maternal blood, breast milk, and the placenta of 6 Japanese women.
    They found that the congeners present were more typical of Kanechlor
    500 than Kanechlor 300, 400, or 600. The results indicated that, as
    the chlorine content of the PCB congeners increased, the correlation
    between the placental content of congeners and those in the maternal
    blood and breast milk also increased. The same was found in laboratory
    animals (Allen & Barsotti, 1976; Masuda et al., 1978b).

    A study on the transfer of PCBs to infants from their mothers was
    carried out in Japan from 1974 to 1976 by Kodama & Ota (1977). When
    the cord blood was considered as the infant blood at birth, the level
    of PCBs in the blood of breast-fed infants rose gradually with
    ingestion of breast milk, exceeded the level in the blood of their
    mothers after 3 months, continued to increase up to the age of 1 year
    and then significantly decreased, 2 years after birth. The PCB
    concentrations in the blood of non-breast-fed infants remained low
    (Table 24).

    6.4  Excretion and elimination

    6.4.1  Following oral dosing

    The excretion of PCBs is, to a large extent, dependent on the
    metabolism of PCBs to form more polar compounds (US EPA, 1987). At
    equilibrium, the elimination of PCBs from all tissues will be
    dependent on the structure-dependent metabolism rates of the
    individual PCB congeners. For example, the biological half-lives in
    the rat range from 1.15 days for 2,2'-dichlorobiphenyl to
    approximately 460 days for 2,4,5,2',4',5',-hexachlorobiphenyl (Tanabe
    et al., 1981; Wyss et al., 1986). Metabolites of the more highly
    chlorinated congeners are eliminated primarily via the faeces (Goto et
    al., 1974; WHO/EURO, 1987).

    When the analysis of faeces is limited to the determination of
    unchanged PCBs, the recovery of the dose administered is incomplete;
    in boars receiving single or repeated doses of Aroclor 1254, not more
    than 16% of the dose was recovered from the faeces and less than 1% in
    the urine (Platonow et al., 1972). Better recoveries have been
    obtained with PCB labelled with radioactive isotopes. Yoshimura et al.
    (1971) found 70% of the activity from a dose of tritium-labelled
    Kanechlor 400 in the faeces and 2% in the urine, over a 4-week period.
    Berlin et al. (1973, 1975) found over 75% of the activity from
    14C-labelled pentachloro- and hexachlorobiphenyls in the faeces and
    less than 2% in the urine; most of the faecal elimination consisted of
    PCB metabolites. Similar results were obtained by Melvs & Brandt
    (1973) with tetrachlorobiphenyls.

    Hashimoto et al. (1976) examined the excretion of 14C-PCB compounds
    given to rats by gavage, at a total dose of 6.35-7.85 mg/kg body
    weight, over a period of 5-50 days. The PCBs studied were
    predominantly tetra- and hexachlorinated isomers. The results
    indicated that 1.9-4.9% of the dose of tetrachlorobiphenyls was
    excreted in the urine, with higher amounts excreted in rats treated
    for longer periods. In rats treated with hexachlorobiphenyls, only
    0.3% of the dose was excreted in urine. About 47-68% of the dose of
    both tetrachloro- and hexachloro-isomers was eliminated in the faeces.

    Table 24.  Level of PCBs in mothers' and babies' blood (average over
               3 years in g/litre)a

    Maternal blood                           4.5          (0.8-15.5)
    Cord blood                               1.1          (nd-5.6)

    Mother's blood (breast-feeding)          2.5          (nd-10.8)
    Babies' blood (3 months old)             3.6          (0.2-10.9)
    Babies' blood (1 year old)               4.7          (0.8-17.7)

    Mother's blood (bottle feeding)          2.7          (0.6-8.7)
    Babies' blood (3 months old)             1.6          (nd-7.6)
    Babies' blood (1 year old)               0.7          (nd-2.1)

    a  From: Kodama & Ota (1977).

    Bleavins et al. (1984) found 22.1% and 1.8% in the faeces and urine,
    respectively, during the first week following dosing of 0.05 mg
    14C-labelled Aroclor 1254 to female European ferrets.

    A biological half-life of about 200 days was recorded in the fat of
    rats after feeding with Aroclor 1254 (Grant et al., 1974). Berlin et
    al. (1975) noted that, in mice dosed with a pentachlorobiphenyl, there
    was an initial rapid elimination from the liver while liver PCB levels
    were high, followed by a slower elimination when most of the PCB was
    located in the fat. The author suggested that the mobilization of PCBs
    from fat, and, therefore, their half-life in the body, depends upon
    their rates of metabolism. Berlin et al. (1973) investigated the
    hypothesis that the ability of a PCB to be readily degraded with a
    half-life of a few days depended on the presence of 2 adjacent
    unsubstituted carbon atoms in the molecule, rather than on the number
    of chlorine atoms, though the presence of such unsubstituted pairs
    depends to a large extent on the degree of chlorination. They came to
    the conclusion that this hypothesis probably applied to unsubstituted
    pairs in the 3,4-position, but that in the 2,3-position, their
    susceptibility to metabolic degradation was influenced more by the
    presence of chlorines in the  o-position of the ring bridge.

    Sprague-Dawley rats, white Swiss mice, and Rhesus monkeys were
    administered a single dose of 14C-2,2'-dichlorobiphenyl, by gavage.
    Within 6 days, mice eliminated a total of approximately 46% of the PCB
    (urine, 20%; faeces, 26%). There was no clear difference between male
    and female mice. In the rat, the total elimination was 51-56% after 9
    days, mainly via the biliary/faecal route. The monkeys had the highest
    elimination rate, a total of 68.6% (urine, 54%; faeces, 14.6%), within
    10 days (Milling et al., 1979).

    Male and female Wistar rats were administered a daily dose of
    14C-2,5,4'-trichlorobiphenyl for 14 days. The animals were killed 5
    days after receiving the last dose. The compound was rapidly
    eliminated primarily with the faeces. Most of the trichlorobiphenyl
    was metabolized (78.5%) and the major metabolites excreted were
    identified as hydroxy-, dichloro-, and conjugated derivatives (Lay et
    al., 1979).

    The elimination of tetrachlorobiphenyl isomers in mice fed diets
    containing a single isomer, at 10 mg/kg diet for 20 days, was studied
    by Mizutani et al. (1977). Biological half-lives for the isomers
    2,3,2',3'-tetrachloro-, 2,4,2',4'-tetrachloro-, 2,5,2',5'-
    tetrachloro-, 3,4,3',4'-tetrachloro-, and 3,5,3',5'-tetrachloro-
    biphenyl, were 0.9, 9.2, 3.4, 0.9, and 2.1 days, respectively.

    Gage & Holm (1976) studied the influence of molecular structure on the
    excretion of 14 PCB congeners in mice. They found that the
    4,4'-dichloro-; 3,3',4',6'-tetrachloro-; 2,3,2',4',6'-pentachloro-;
    and 2,3,4,2',4',5'-hexachloro-isomers were eliminated most rapidly.
    These compounds had at least one pair of unsubstituted  ortho-meta,
    vicinal carbon atoms, a configuration thought to be important for
    rapid metabolism and excretion. The most slowly eliminated compounds
    were 2,4,5,2',4',5'- and 2,3,4,2',4',5'-hexachlorobiphenyl.

    2,4,5,2',4',5'-Hexachlorobiphenyl was the PCB congener found in the
    highest concentration in human adipose tissue, while
    2,4,6,2',4',6'-hexachlorobiphenyl was not detected (Jensen &
    Sundstrm, 1974a). As both of these compounds are found in commercial
    PCB mixtures and in the environment, the presence of the
    2,4,5,2',4',5'-hexachlorobiphenyl in adipose tissue appears to be
    related to resistance to metabolism (US EPA, 1987). That this congener
    is not, or is only minimally, metabolized is also indicated by the
    finding that the blood concentration of this congener decreased only
    10% over 300-500 days (Chen et al., 1982) and by the results of
     in vitro metabolism studies with human liver microsomes (Schnellman
    et al., 1984a,b).

    Felt et al. (1977) examined the elimination of 14C-2,5,4'-trichloro-
    biphenyl in rhesus monkeys. The monkeys were fed 550 mg of the
    compound in fruit, daily, for 84 days. On the basis of total excretion
    and recovered radioactivity, the half-life was found to be 4.5-4.8

    Male and female Sprague-Dawley rats were administered 14C-labelled
    2,4,6,2',4'-pentachlorobiphenyl by gavage and the urine and faeces
    were collected. After 8 days, the animals were killed. The elimination
    of the PCBs followed a bi-exponential rate expression with a-phase
    half-lives of 0.90 and 0.95 days and b-phase half-lives of 4.2 and 3.8
    days, for males and females, respectively (Felt et al., 1979).

    6.4.2  Following parenteral dosing

    The results of injection studies indicate that PCBs can be excreted
    unmetabolized into the gastrointestinal tract. Yoshimura & Yamamoto
    (1975) recovered unchanged tetrachlorobiphenyl from the duodenal
    contents of rats injected intravenously with tetrachlorobiphenyl.
    Daily excretion for 4 days ranged from 0.5 to 0.8% of the total
    dose/day. Goto et al. (1974) found that 4.7-23.2% of injected PCBs
    were excreted unchanged into the gastrointestinal tract by day 10
    after dosing, with the excretion of a penta-isomer greater than the
    excretion of di-, tri-, or tetra-isomers.

    Adult male Sprague-Dawley rats received doses of 4 symmetrical
    hexachlorobiphenyl 14C-isomers, i.e., 2,3,5,2',3',5'-hexachloro-,
    2,3,6,2',3',6'-hexachloro-, 2,4,5,2',4',5'-hexachloro-, and
    2,4,6,2',4',6'-hexachlorobiphenyl, by intravenous injection. Most of
    the radioactivity was eliminated in the faeces with less than 1% found
    in the urine. The metabolites showed evidence of dechlorination,
    chlorine shifts, and possible metabolism by direct insertion of a
    hydroxyl group. There was also evidence supporting the intermediate
    step of an arene-oxide as a predominant mechanism of PCB metabolism
    (Kato et al., 1980).

    The disposition of 2 symmetrical 14C-labelled 2,3,6,2',3',6'-
    hexachloro- and 2,4,5,2',4',5'-hexachlorobiphenyl was studied in
    24-month-old, male, Sprague-Dawley rats, after iv treatment. More than
    50% of the 2,3,6,2',3',6'-hexachlorobiphenyl was metabolized and
    excreted via the bile into the faeces within 2 days, and only 2% was
    excreted in urine. More than 90% was eliminated as metabolites. In
    contrast, 2,4,5,2',4',5'-hexachlorobiphenyl was redistributed from the
    liver, muscle, and skin to the adipose tissue, where it accumulated
    without being metabolized. Only 2% of the total dose was eliminated,
    primarily in the faeces, within 21 days. In 2- to 3-month-old rats,
    the general pattern of disposition of these hexachlorobiphenyls did
    not change with age; however, there were differences in the rates of
    elimination and in the tissue levels. There was enhanced metabolic
    retention in the muscle, skin, and adipose tissue of older rats, which
    suggested an age-related decrease in tissue clearance. The larger
    volume of adipose tissue could not explain this observation. In
    general, there were few changes in decay rates from tissues or in
    biliary excretion, so age had a greater effect on the disposition of
    the "persistent" 2,4,5,2',4',5'-hexachlorobiphenyl than on the
    metabolizable 2,3,6,2',3',6'-hexachlorobiphenyl (Birnbaum, 1983).

    Ethane exhalation was increased in male Sprague-Dawley rats, 30 days
    after a single ip injection of Aroclor 1254 (500 mg/kg body weight).
    Before day 30, there was no increase in ethane production. Parallel
    increases in hepatic malondialdehyde levels were found. A single ip
    injection of 3,4,3',4'-tetrachloro-, 2,3,4,5,4'-pentachloro, and
    2,4,5,2',4',5'-hexachlorobiphenyl (300 mol/kg) also increased (after
    30 days) the production of malondialdehyde and ethane, indicators of
     in vivo lipid peroxidation. These effects were not reflected in
    increased diene conjugation (Dogra et al., 1988).

    Sipes et al. (1980, 1982a,b) studied the distribution, metabolism, and
    excretion of 14C-labelled 4,4-dichloro-, 2,4,5,2'4'5'-hexachloro-, or
    2,3,6,2',3',6'-hexachlorobiphenyl in beagle dogs and cynomolgus
    monkeys, after a single intravenous dose. The elimination of the test
    substances from the blood of both species was shown to be biphasic.
    The results for dichlorobiphenyl showed that the dog eliminated 50% of
    the dose (urine, 7%; faeces, 43%) within 24 h, while the remainder was

    found mainly in the adipose tissue. By 5 days, 90% had been
    eliminated. The monkey eliminated less than 15% of the dose within
    24 h, with less than 1% in the faeces. The remainder was found in the
    adipose tissue. Within 28 days, 59% of the dose had been eliminated,
    chiefly in the urine. Biliary excretion after 24 h was shown to be 33%
    in the dog and only 0.4% in the monkey.

    The data for 2,4,5,2',4',5'-hexachlorobiphenyl showed that the dog
    eliminated 66% (urine, 3%; faeces, 63%) within 3 days; the monkey
    eliminated 18% of the dose (of which 17% was in the faeces), 90 days
    following administration. The remainder was found in the adipose
    tissue. In the studies with 2,3,6,2',3',6'-hexachlorobiphenyl, the dog
    eliminated 52% of the dose within 24 h (urine, 11%; faeces, 41%) and
    70% in 3 days. The monkey eliminated 19% during the first 24 h,
    divided equally between urine and faeces. By 15 days, 61% had been
    eliminated, primarily in the faeces. The 24-h biliary excretion was
    26% and 2.4% in the dog and the monkey, respectively.

    6.4.3  Humans

    Chen et al. (1982, 1985) studied the presence of PCBs in the blood of
    human beings, in the Province of Taiwan, after they had consumed
    rice-bran oil contaminated with Kanechlor 500 and PCDFs. Blood samples
    from 17 patients were examined, with 2-3 samples taken from each
    patient, 2-17 months apart. The results indicated that the
    tetrachloro- and some pentachloro- isomers tended to be eliminated
    more rapidly than the other pentachloro- and the hexachloro- and
    heptachloro- isomers. Half-lives for the 2,4,5,2',4'- and 2,3,4,3',4'-
    pentachloro- isomers in the blood were 9.8 and 8.7 months,
    respectively. Two adjacent unsubstituted carbon atoms at the  meta,
     para positions facilitated metabolism and the subsequent elimination
    from the blood. PCBs containing adjacent unsubstituted carbon atoms at
    the  ortho and  meta positions of the biphenyl ring are eliminated
    very slowly and will accumulate.

    Buhler et al. (1988) administered a uniformly 13C-labelled PCB
    mixture similar to Aroclor 1254 to a volunteer. A single dose of
    329 g/kg body weight was ingested; blood samples taken over a period
    of 260 days were analysed for 13C- and 12C-PCBs using GC/MS and
    GC/ECD. Elimination of the isomers followed a first order kinetics.
    The half-lives for the isomers 2,3,4,2',4',5'-hexachlorobiphenyl,
    2,4,5,2',4',5'-hexachlorobiphenyl, and 2,3,4,5,2',4',5'-hepta-
    chlorobiphenyl were 321, 338, and 124 days, respectively.

    6.4.4  Elimination via milk (animals)

    Vodicnik (1986) studied the disposition of 14C-2,4,2',4'-
    tetrachlorobiphenyl (150 mg/kg body weight administered
    intraperitoneally) as a function of non-pregnant body weight in
    virgin, late pregnant, and early post partum ICR mice and their
    offspring. The highest concentrations were observed in adipose tissue
    and the mammary glands, regardless of reproductive state. The
    concentrations of the tetrachlorobiphenyl equivalents in the tissues
    differed among the 3 groups, possibly because of the alterations in
    lipid deposition/mobilization associated with pregnancy and lactation.
    Approximately 20% of 14C-activity was eliminated from the carcass of
    virgin mice, 4 days after administration, but no decrease was seen in
    late-pregnant animals. Minimal transplacental transfer of
    14C-activity occurred (approximately 1%), but the tetrachlorobiphenyl
    was rapidly eliminated in breast milk to nursing offspring. Ninety per
    cent of the total-carcass 14C-activity was eliminated from lactating
    mice over a 4-day period, approximately 75% of which could be
    accounted for in neonatal carcasses.

    Saschenbrecker et al. (1972) found that, after oral administration of
    doses of Aroclor 1254 of 10 or 100 mg/kg to cows, 6.27 and
    74.5 mg/litre, respectively, appeared in the milk after 24 h. These
    levels were reduced to less than one-half within 3 days, but traces
    still remained at 50 days. Cows receiving 200 mg/day of Aroclor 1254
    reached a steady state concentration of 61 mg/kg in the milk fat and
    42 mg/kg in the body fat, after 10 days (Fries et al., 1973).

    The "carry-over factor" from animal feed into the cow's milk showed
    that the lower (tri-, tetra-, and penta-) chlorinated biphenyls have a
    lower carry-over factor than the higher (hexa- and hepta-) chlorinated
    biphenyls. Thus, it is the latter that are particularly concentrated
    in cow's milk fat. From studies in the Federal Republic of Germany, it
    was found that the major congeners in cow's milk were numbers 138,
    153, and 180 (DFG, 1988).  Elimination via breast milk

    The composition of common commercial PCB mixtures clearly differs from
    the composition of the PCB contents of human fat or human breast milk,
    because of the preferential elimination of certain PCB congeners
    containing 3 or 4  ortho substituents and the retention of PCBs with
    1 or 2  ortho substituents (Kuroki & Masuda, 1977; Watanabe et al.,
    1979; Yakushiji et al., 1979).

    The major PCB components (and average relative concentrations) that
    have been identified in breast milk in the Osaka area in Japan
    include: 2,4,4'-trichlorobiphenyl (8.4%); 2,5,2',5'-tetrachloro-
    biphenyl (2.0%); 2,4,5,4'-tetrachlorobiphenyl (19%); 2,4,5,2',5'-
    pentachlorobiphenyl (2.8%); 2,4,5,3',4'-pentachlorobiphenyl (11.8%);
    2,4,5,2',4',5'-hexachlorobiphenyl (15.5%); 2,3,4,2',4',5'-
    hexachlorobiphenyl (15.8%); 2,3,4,3',4',5'-hexachlorobiphenyl (2.3%);
    2,3,4,6,2',4',5'-heptachlorobiphenyl (1.6%); 2,3,5,6,2',4',5'-
    heptachlorobiphenyl (3.2%). These PCB-congeners constituted at least
    95% of the PCBs in the breast milk of the women examined in Osaka.

    In recent studies, the contents of PCBs in human milk and maternal
    blood were compared for US citizens (Bush et al., 1984, 1985). Eight
    individual PCB congeners comprised 52% of the total PCB residues in
    the milk and 48.5% in the blood. The mean concentrations for total
    PCBs were 26.5 g/kg for whole milk and 3.5 g/kg for blood. The
    percentages of the different congeners are given in Table 25.

    6.5  Metabolic transformation

    6.5.1  PCBs

    The metabolism of PCBs has been investigated in numerous studies on
    animals and reviewed by Drill et al. (1981) and the US EPA (1987). The
    PCBs were usually administered by the oral or parenteral route.

    Phenolic products are the major PCB metabolites, though
    sulfur-containing metabolites, trans-dihydrodiols, polyhydroxylated
    PCBs, and methyl ether derivatives have also been identified. Although
    the effects of the chlorine substitution pattern on sites of oxidation
    have not been studied systematically, US EPA (1987) suggested the

    *   hydroxylation is favoured at the  para position in the least
        chlorinated phenyl ring, unless this site is sterically hindered
        (i.e., 3,5-dichloro-substitution);

    *   in the lower chlorinated biphenyls the  para position of both
        biphenyl rings and carbon atoms that are  para to the chloro
        substituent are all readily hydroxylated (Sparling et al., 1980);

    *   the availability of 2 vicinal unsubstituted carbon atoms
        (particularly C5 and C4 in the biphenyl nucleus) also facilitates
        the oxidative metabolism of the PCB substrate, but is not a
        necessary requirement for metabolism;

        Table 25.  Concentrations of most abundant PCB congeners present in whole breast milk and maternal blooda

    Congener                                 Milk                                 Maternal blood                  Ratio
                                             (40 samples)                         (101 samples)                   milk/blood

                                             g/litre          % of               g/litre     % of
                                                               total PCBs                      total PCBs

    2,4,5,2',4',5'-hexachlorobiphenyl          3.2              12                 0.31           8.8               10
    2,3,5,6,2',3',6-heptachlorobiphenyl        2.5               9.4               0.27           8.0                9.2
    2,4,5,2',3',4'-hexachlorobiphenyl          2.1               7.8               0.58          17                  3.5
    2,5,3'4'-tetrachlorobiphenyl               1.7               6.6               0.01           -                500
    2,3,4,5,2'4'5'-heptachlorobiphenyl         1.2               4.5               0.03           3.7                9.4
    2,3,4,5,3',4'-hexachlorobiphenyl           1.0               4.0               0.01           -                125
    2,4,5,2',4'-pentachlorobiphenyl            1.1               4.0               0.12           3.4                8.9
    2,3,4,3',4'-pentachlorobiphenyl            0.97              3.7               0.25           7.6                3.8

    Total PCBs                                26.5               -                 3.5            -                  7.5

    a  Modified from: Bush et al. (1985).

    *   as the rate of chlorination increases on both phenyl rings, the
        rate of metabolism decreases;

    *   the metabolism of specific PCB isomers by different species can
        result in considerable variations in metabolic pattern.

    Kannan et al. (1989) studied the possible involvement of frontier
     (pi) electrons in the metabolism of polychlorinated biphenyls. The
    electron density, at each carbon atom, of the highest occupied  pi
    orbital of 13 PCB molecules was calculated and the result was compared
    with their  in vitro and/or  in vivo metabolism. It was found that:

    *   the carbon position at which the frontier electron density was the
        highest was most readily hydroxylated or sulfonated;

    *   if the carbon with the highest frontier  (pi) electrons was
        occupied by chlorine, either a replacement occurred or the carbon
        with the next highest electron density was activated for

    *   because of steric hindrance,  "ortho" carbons were least
        preferred for such reactions, in spite of possessing favourable
        electron density;

    *   this was applicable to both phenobarbital (PB)-type and
        3-methylcholanthrene (3-MC)-type PCB inducers.

    The authors suggested that frontier  (pi) electron density could be
    an easy guide for understanding the metabolic products of persistent
    and toxic environmental pollutants  in vitro and  in vivo, and for
    understanding their environmental fate.

    There appears to be little metabolism of PCBs with 6 or more chlorine
    substituents (Matthews & Anderson, 1975b). When between 2 and 5
    chlorine substituent PCBs are metabolized, the metabolic products are
    primarily hydroxylated compounds, frequently found as glucuronide
    conjugates (hydroxymethoxy derivatives) and partially dechlorinated
    metabolites. In some cases, smaller amounts of dihydrohydroxy
    compounds and related substances are also found.

    The parent compound is also eliminated in various quantities in
    faeces, hair, and maternal milk, but very little unmetabolized
    compound is excreted in the urine. This pattern is not unusual for
    lipophilic xenobiotics.

    PCB metabolism has been examined in primates (monkeys) by Greb et al.
    (1975), Hsu et al. (1975a,b), and Allen & Norback (1976); in ungulates
    (cows, pigs, and goats) by Platanow & Chen (1973), Safe et al. (1975),
    and Gardner et al. (1976); in rats by Grant et al. (1971a), Hutzinger
    et al. (1972), Yoshimura et al. (1973), Goto et al. (1973, 1974,
    1975), Safe et al. (1974), Matthews & Anderson (1975b), van Miller et
    al. (1975), Sundstrm & Jansson (1975), Sundstrm et al. (1976a), Lay
    et al. (1975, 1979), Chen et al. (1976), Kamal et al. (1976), and
    Norback et al. (1976); in mice by Berlin et al. (1973), Yamamoto &
    Yoshimura (1973), and Sundstrm & Jansson (1975); in rabbits by Grant
    et al. (1971b), Hutzinger et al. (1974), Sundstrm & Wachmeister
    (1975), and Sundstrm et al. (1976b); in pigeons, and quails by Koeman
    et al. (1969), Hutzinger et al. (1972), Bailey & Bunyan (1972), and
    Sundstrm & Jansson (1975); and in trout by Hutzinger et al. (1972).

    The different metabolic products formed from pure isomers in these
    various species have been catalogued in an NAS report (1979) and in a
    review by Sundstrm et al. (1976a). Neither of these reports is
    complete, but, together, they cover most of the studies up to 1979.

    In the rat, monochloro-, dichloro-, trichloro-, tetrachloro-,
    pentachloro-, and at least one hexachlorobiphenyl, yielded at least
    one hydroxylated metabolite. Some isomers produced as many as 5
    different hydroxylated metabolites including both mono- and dihydroxy-
    derivatives. Most of the hexachloro-, octachloro-, and
    decachlorobiphenyls did not yield detectable levels of hydroxylated

    Similar hydroxylated derivatives were also produced in other species,
    but the ability to metabolize PCBs is not absolutely uniform in all
    species. In the rabbit, dichloro-, tetrachloro-, and
    hexachlorobiphenyls were metabolized, while further down the
    phylogenetic scale, the pigeon only metabolized monochloro- and
    dichlorobiphenyls and the trout failed to metabolize any of the
    chlorinated biphenyls tested. Table 26 shows the PCBs tested in
    different species and indicates whether or not the organism was able
    to metabolize the compound. Although different species may metabolize
    a given isomer, the metabolic products are not necessarily identical.
    An example of this is found in the simple 4,4'-dichlorobiphenyl which
    is metabolized by the rat, rabbit, and goat, but does not give
    identical products in these species; all 3 species produce
    4,4'-dichloro-3-hydroxybiphenyl as a metabolite, but, in addition, the
    rat produces 4,4'-dichloro-, 2,3-dihydroxybiphenyl, and the goat
    produces 3,4'-dichloro-4-hydroxybiphenyl as a metabolite.

    However, a product such as the 3,4'-dichloro-4-hydroxybiphenyl found
    in the goat involves a chlorine shift, which may be indicative of a
    more toxic intermediate.

    Many different pathways of metabolism have been described as
    summarized in Fig. 5 (Safe, 1984; WHO/EURO, 1987).

    These pathways include hydroxylation, and conjugation with thiols and
    other water-soluble derivatives. The most important pathway seems to
    be through hydroxylation and subsequent conjugation. Rats and mice
    that were exposed to dichloro-, tetrachloro-, or pentachlorobiphenyls
    by intraperitoneal injection or diet, eliminated metabolites as
    glutathione conjugates and other sulfur-containing compounds (Kurachi,
    1983; Kurachi & Mio, 1983). Mammalian metabolism of many individual
    PCBs may proceed via oxide intermediates, which have not been
    isolated, but are presumed to be precursors of some of the major
    metabolites identified. One type of metabolite is the methylsulfone
    PCB metabolite that has been identified in environmental samples by
    Jansson et al. (1975) and WHO/EURO (1987), and in human milk by
    Yoshida & Nakamura (1979).

    The formation of xenobiotic thioether derivatives, including
    glutathione, cysteinylglycine, cysteine, and  N-acetylcysteine
    (mercapturic acid) conjugates, is generally considered a pathway for
    the detoxification of reactive intermediates. Mio & Sumino (1985)
    detected methylsulfonyl metabolites, by using GC/MS/COM, from the
    adipose tissues of mice treated with Kanechlor 300, 400, 500, or
    2,5,2',5'-tetrachlorobiphenyl. Metabolites were detected in the faeces
    of mice treated with 2,5,2',5'-tetrachlorobiphenyl, e.g., 6
    sulfur-containing and 5 non-sulfur-containing metabolites. The
    elimination rates for one week were 1.7% and 43%, respectively. The
    methylsulfonyl metabolites accumulated in the liver, adipose tissue,
    and lungs. Mio & Sumino (1985) proposed the methylsulfonyl metabolic
    pathway of 2,5,2',5'-tetrachlorobiphenyl.

    Klasson-Wehler et al. (1987) administered a single dose of 2,3,6,4'-
    tetrachlorobiphenyl to 3 groups of 5 female C57B1 mice at 0, 10, or
    100 mg/kg body weight. 35S-cysteine was administered by ip injections
    4 times at 12-h intervals. The animals were sacrificed and the organs
    analysed on day 12. Methyl [35S]sulfonyl-tetrachlorobiphenyl was
    found in the lungs, kidneys, and fat of the treated mice, as well as
    minor amounts of tetrachlorobiphenyl and traces of

    The formation of serial methylsulfonyl metabolites can be summarized
    as follows: the glutathione conjugate is converted, by cleavage, to a
    cysteine or thiol conjugate and translocated into the liver. The thiol
    conjugate from the cysteine moiety is transmethylated by
    thiol- S-methyltransferase and is oxygenated by cytochromes P-450 and
    P-448 oxidase or is glucuronidated by UDP-glucuronyl-transferase in
    the liver, resulting in methylsulfonyl derivatives.

        Table 26.  Metabolism of various PCBs in different organisms

    Compound                        Species

    Chlorobiphenyl                  Trout    Pigeon    Mouse    Rat   Rabbit   Monkey

    4-mono-                         -        +                  +     +

    4,4'-di-                        -        +                  +     +
    2,2'                                                        +
    2,4'                                                        +              +

    2,5,2'-tri                                                                 +

    2,4,2',4'-tetra-                                            +
    2,5,2',5'                       -        -                  +     +        +
    3,4,3',4'                                                   -

    2,3,4,5,6,-penta-                                           +
    2,4,5,2',5'                                        +        +
    2,4,6,2',4'                                                 +
    2,4,6,2',6'                                                 +
    2,4,6,3',5'                                                 +
    2,3,4,6,4'                                                  +
    2,3,4,3',4'                                                 -

    2,4,5,2',4',5',-hexa-           -        -                  ()a  +
    2,4,6,2',4',6'                                              ()a

    2,3,5,6,2',3',5',6'-octa-                          -

    2,3,4,5,6,2',3',4',5',6'-deca                      -

    a  These compounds were reported by Sundstrm et al. (1976a) as failing to produce
       hydroxylated derivatives, but were positive in an IARC report referred to in an
       NAS report (NAS, 1979).
       + = Compound is metabolized; - = Compound not metabolized: (blank) not tested.

    FIGURE 5

    The occurrence of  trans-dihydrodiol metabolites suggests that the
    metabolism of PCBs proceeds through the formation of arene oxide
    intermediates (US EPA, 1987). Arene oxides are potential electrophiles
    that have been implicated in cellular necrosis, mutagenicity, and
    carcinogenicity (Safe et al., 1975; Sundstrm et al., 1976a).

    While the arene oxide pathway is important in carcinogenic
    considerations, it may not be the primary pathway for the metabolism
    of PCBs, in most cases. So far, most discussions about the metabolism
    of PCB isomers have focused on the position and number of the chlorine

    The metabolic products of dichloro-, trichloro-, tetrachloro-,
    pentachloro-, and hexachlorobiphenyl appear to reflect direct
    hydroxylation at the  meta and/or  para positions, relative to the
    position of the phenyl-phenyl bond. In a few instances, a methoxy
    group is found instead of a second hydroxyl. This direct mechanism
    appears to operate, therefore, irrespective of the degree of
    chlorination, and, for the most part, irrespective of the position of
    the chlorine substituents. In the rabbit, some exceptions have been
    found that involve: a chlorine shift and the removal of a chlorine, in
    the case of the 4,4'-dichlorobiphenyl, the formation of a dihydrodiol
    at the  meta and  para positions, in the case of 2,5,2',5'-tetra-
    chlorobiphenyl, and the removal of a chlorine from one of the rings,
    in the case of 2,4,5,2',4',5'-hexachlorobiphenyl (Sundstrm et al.,

    Studies carried out by Matthew & Anderson (1975b) and Tuey & Matthews
    (1977) showed that monochloro- and dichlorobiphenyls were rapidly
    metabolized and excreted and that pentachloro- and hexachlorobiphenyls
    were poorly metabolized and retained longer in the adipose tissue and
    skin. The situation for the tetrachlorobiphenyls is more complicated.

    The following analysis is based on the data on metabolites identified
    and reported in the review by Sundstrm et al. (1976a).

    6.5.2  Dichlorobiphenyls

    Consideration of the various dichloro- isomers shows that, when
    chlorines are only on one ring, hydroxylation occurs on the
    nonchlorinated ring. Single hydroxylation occurs  para to the
    phenyl-phenyl bond; if another hydroxylation occurs, it is always
     meta to the phenyl-phenyl bond. This holds true for the 3 different
    isomers tested: 2,3-; 2,4-; and 3,4-dichlorobiphenyls. When the
    dichloro-compounds are symmetrically chlorinated on each ring, as in
    2,2'-; 3,3'-; and 4,4'-dichloro compounds, the same pattern applies
    generally, but with a variation on the theme and an exception in the
    case of 4,4'-dichlorobiphenyl. In the case of 2,2'- and 3,3'-dichloro

    compounds, monohydroxylation occurs  meta and  para to the
    phenyl-phenyl ring, respectively. In both cases, double hydroxylation
    involves both  meta and  para positions on the same ring, (that is,
     meta and  para to the phenyl-phenyl bond). In the case of
    4,4'-dichlorobiphenyl, there appears to be a difference in the rat.
    The monohydroxy- derivative is  meta, but the dihydroxy derivative is
     ortho and  meta to the phenyl-phenyl bond. Not only is the rat
    metabolism of the 4,4'-compound an exception, but the rabbit also
    shows an unusual response to this compound. In the rabbit, the
    monohydroxy-derivative is the same  meta hydroxy found in the rat,
    but, instead of a dihydroxy- compound, the rabbit produces a chlorine
    shift and a single hydroxy group in the  para position as well as
    dechlorination and hydroxylation in the  para position. These latter
    products have been considered as characteristic of the arene oxide
    intermediate pathway. Nevertheless, among the 6 different dichloro-
    isomers examined, all produced a monohydroxy- derivative, either  meta
    or  para to the phenyl-phenyl bond, and all but the 4,4' produced
    dihydroxy- derivatives were  meta and  para on the same ring to the
    phenyl-phenyl bond.

    The absence of a substitution at 4,4'- with vicinal unsubstituted
    positions cannot be correlated with rapid metabolism, since this
    property is also shared by both a rapid and slowly metabolized isomer.
    This is in direct contradiction to the often repeated statement "the
    presence of at least two adjacent unsubstituted carbons, particularly
    in positions 3,4-, or 5- or 3',4'- or 5'- is required for rapid
    metabolism of chloro-biphenyl" (Jensen & Sundstrm, 1974b; Berlin et
    al., 1975; Safe et al., 1975; Matthews & Anderson, 1976; NIOSH, 1977;
    Matthews & Tuey, 1980).

    6.5.3  Tetrachlorobiphenyls

    Examination of the 2 different tetrachloro- isomers, 2,3,5,6,-
    tetrachloro- and 2,5,2',5',-tetrachlorobiphenyl, showed that, in the
    case of the molecule with all 4 chlorines on one ring, the products
    were monohydroxy- derivatives  meta or  para, dihydroxy- derivatives
     meta and  para, and a  para hydroxy- plus a  meta methoxy- group
    or a  meta hydroxy- and a  para methoxy- group, all on the
    unsubstituted ring. The symmetrical 2,5,2',5'-tetrachlorobiphenyl in
    the rat gave the  meta hydroxy-, but in the rabbit a  para hydroxy-,
    and also in the rabbit a dihydro-dihydroxy-, on the  meta and  para
    positions. The asymmetric 2,4,3',4'-tetrachlorobiphenyl gave
    monohydroxy-derivatives, both in the  meta position, either in the
    three or five position. It seems that an alternative enzyme pathway is
    available in the rabbit.

    The level of retention was highest for 2,4,2'4'-tetrachlorobiphenyl
    descending in the following order; 2,5,2',5'-, 3,5,3',5'-, 3,4,3',4'-,
    2,3,2',3'-, and 2,6,2',6'-tetrachlorobiphenyl. Since no PCBs were
    detected in the liver, and only small amounts of 2,3,2',3'- and
    3,4,3',4'-tetrachlorobiphenyls in the carcass, it can be concluded
    that these 2 isomers were readily metabolized and excreted. Both
    compounds have unsubstituted vicinal positions. However, the compounds
    slowest to be metabolized were 2,4,2',4'- and 2,5,2',5'-tetra-
    chlorobiphenyls, which also have unsubstituted vicinal positions.
    Metabolic restriction cannot be entirely attributed to substitution at
    the 4,4'- positions (Kato et al., 1980), since this also occurred in
    3,4,3',4'-tetrachlorobiphenyl, which was removed relatively rapidly.

    The excretion of the monohydroxy metabolites of 3,4,3',4'-tetra-
    chlorobiphenyl and 2,4,3',4'-tetrachlorobiphenyl (orally administered)
    in rats has been demonstrated by Yoshimura et al. (1973); Yamamoto &
    Yoshimura (1973); Yoshimura & Yamamoto (1975); and Yoshimura et al.
    (1974). They demonstrated that the metabolites of the first isomer
    were 2-hydroxy- or 5-hydroxy- compounds, while the metabolites of the
    second isomer were 5-hydroxy- and 3-hydroxy- compounds. All hydroxy
    metabolites were excreted non-conjugated via the bile and no parent
    isomers were found in the bile. Yoshimura & Yamamoto (1975) found that
    unchanged 2,4,3',4'-tetrachlorobiphenyl was excreted through the
    intestine, when it was intravenously injected in rats with the bile
    duct ligated, while no metabolite of this isomer was excreted by this

    The results with the 2,5,2',5'- molecule are probably related to an
    arene oxide pathway. Direct evidence for this was reported by Forgue
    et al. (1980), who showed that 3,3,3,-trichloropropene-1,2-oxide,
    which is an inhibitor of epoxide hydrase, blocked the formation of the
    suspected arene oxide metabolites. The arene oxide mechanism
    supposedly operates in rabbits and monkeys for 2,5,2',5'-tetrachloro-
    biphenyl and, possibly, also in rats for 2,4,5,2',4',5'-hexachloro-
    biphenyl. Isomers that may utilize the arene oxide pathway, to some
    extent, are: 4,4'-dichloro-; 2,5,2',5'-tetrachloro-; and

    6.5.4  Hexachlorobiphenyls and higher chlorinated compounds

    The symmetrical hexachlorobiphenyls were used in a study by Matthews &
    Tuey (1980), in which Sprague-Dawley rats were injected intravenously
    with the PCBs and killed at increasing time intervals from 15 min to
    42 days. The 2,3,6,2',3',6'- isomer was rapidly metabolized and
    excreted compared with the other isomers, which were slowly
    metabolized and excreted with much longer half-lives. The results
    indicated that the metabolism of hexachlorobiphenyls is slow, when the
    position of the chlorine atoms is such that arene oxide formation is

    2,3,6,2',3',6'-Hexachlorobiphenyl produces only one metabolite in the
    rat: 2,3,6,2',3',6'-hexachloro-4'-hydroxybiphenyl, which is believed
    to be the result of arene oxide formation. All commercial mixtures of
    PCBs will contain congeners that could be metabolized via the arene
    oxide pathway. However, it does not seem to be the major pathway of
    metabolism for most of the components of the commercial products,
    since most of the higher congeners will not have vicinal unsubstituted

    The metabolic data on individual isomers shows that, at least up to
    hexachloro- compounds, ordinary hydroxylation can take place. It is
    reasonable to consider that it is not only as a consequence of poor
    metabolism that pentachloro- and hexachloro- compounds are persistent
    in the tissues, but rather that they are not metabolized as readily,
    because they are sequestered from tissues in which the bulk of the
    metabolism takes place. In support of this position, it has been shown
    by Matthews & Anderson (1975b) that, when animals are caused to lose a
    substantial portion of body weight, the stored higher chlorinated
    compounds can, indeed, be metabolized. Octachloro- and
    decachlorobiphenyls would not be expected to be easily metabolized,
    simply because there are few or no sites for hydroxylation to take
    place. It was found (Vodicnick & Lech, 1980; Vodicnick et al., 1980)
    that almost the entire body burden of 2,4,5,2',4',5'-hexa-
    chlorobiphenyl was removed from mothers given this PCB and that it was
    transferred to their offspring via the nursing mother's milk. In mice,
    the preferential distribution of this PCB in milk reflects the high
    fat content of mouse milk.

    Virgin, female, Sprague-Dawley mice, injected ip with 50 or 100 mg
    [14C] 2,4,5,2',4', 5' -hexachlorobiphenyl/kg body weight in corn oil
    for 2 weeks, prior to mating, eliminated virtually their entire body
    burden of the compound through milk during one lactation cycle
    (Gallenberg & Vodicnick, 1987).

    Storage is caused by lack of metabolism and also implies that the
    availability of adjacent unsubstituted carbons is the determinant for
    metabolism. Direct hydroxylation reactions do not require
    unsubstituted adjacent carbons. Rapid storage in fat is the
    rate-limiting factor in the removal of most PCBs, with the exception
    of isomers that might be very rapidly metabolized by arene oxide
    formation, such as 2,3,6,2',3',6'-hexachlorobiphenyl.

    Matthews & Anderson (1975b) extended their study to include a fasting
    period to reduce the weight of the test rats and showed that severe
    fasting mobilized stored PCBs and brought them into the metabolic

    6.5.5  Retention and turnover

    Mizutani et al. (1977) studied the pharmacokinetic behaviour of 6
    different tetrachlorobiphenyls. They administered mice the 6 isomers,
    at 10 mg/kg body weight, for 20 days. The isomer concentrations in the
    liver and the remainder of the carcass were determined at various
    times during the recovery period. They found that the accumulated body
    burden was a function of both storage ratio and biological half-life.

    The results of Mizutani et al. (1977) suggest that the correlations
    claimed between the position of the chlorine substituents and storage
    or metabolic activity (Kato et al., 1980; Matthews & Tuey, 1980) are
    not simply explained.

    The hypothesis that the position of the chlorine atoms alone
    determines the rates of metabolism, accumulation, and excretion does
    not appear to be entirely supported. This idea has been used to
    support the notion that PCBs with unsubstituted vicinal carbon atoms
    favour metabolism by arene oxide formation.

    6.5.6  Appraisal

    The results of most studies suggest that PCBs are absorbed by the
    organ systems (gastrointestinal tract, lung, and liver), representing
    the likely routes of entry into the body. This is particularly true
    for the gastrointestinal tract where absorption is rapid. PCBs, once
    absorbed, are usually distributed in a biphasic manner and are rapidly
    cleared from the blood and accumulated in the liver and adipose
    tissue, or they can be metabolized in the liver, to form metabolites
    that are excreted in the urine and bile. In some studies on humans,
    the skin, an organ rich in adipose tissue, had a high PCB content,
    whereas the brain content was low. This distribution can also include
    the fetus and human milk, an extension of the adipose tissue system in
    the body. Mobilization of PCBs from fat appears to depend on their
    rates of metabolism. Metabolic pathways include hydroxylation, and
    conjugation with thiols and other water-soluble derivatives, some of
    which can involve reactive intermediates, such as the arene oxides.
    The most important pathway seems to be hydroxylation and subsequent
    conjugation. This pathway is facilitated by the presence of at least
    one pair of unoccupied vicinal carbon atoms in the PCB structure.
    Persistence in tissue is not correlated with high toxicity.
    Differences in toxicities among PCBs may be associated with specific
    metabolites and/or their associated intermediates.


    7.1  Toxicity for microorganisms

    7.1.1 Freshwater microorganisms

    Zullei & Benecke (1978) used the motility of filamentous blue-green
    algae (Cyanophyceae) of the genus  Phormium as a bioassay for the
    rapid determination of the toxicity of various compounds. They tested
    the relative toxicity of purified, specific chlorobiphenyls and of PCB
    commercial mixtures (Clophen). Inhibition of motility was greatest in
    the presence of chlorobiphenyls of low chlorination. All mono- and
    dichlorinated isomers were inhibitory at the test concentration of
    100 g per test spot of algae. Tetra- and hexachlorobiphenyls did not
    have any effects; the effects of trichlorobiphenyl isomers varied, the
    2,5,2' isomer not producing any effects and the 2,4,4' and 3,4,4'
    isomers being inhibitory. Tests using the commercial mixtures
    confirmed the greater toxicity of low chlorination levels; Clophen A30
    was more toxic than Clophen A60, though the presence of more than
    expected, low-chlorinated compounds in the Clophen A60 reduced the
    difference in toxicity.

    Cultures of the green alga  Chlorella pyrenoidosa were incubated with
    1 mg/litre of Aroclors 1242, 1254, and 1268 (Hawes et al., 1976a).
    The initial culture was 5 days old with a cell density of 27 
    106 cells/ml. After 8 h of incubation with the PCBs, cell densities
    in the cultures were 64% lower than controls for Aroclor 1242, 45%
    lower for Aroclor 1254, and 36% lower for Aroclor 1268. As the study
    progressed, the cell densities in the cultures improved relative to
    the controls; cell density in the Aroclor 1254 culture was equal to
    that in the control by 129 h and the density in the Aroclor 1268
    culture was equal to the control by 59 h. Although the density in the
    Aroclor 1242 culture remained much lower than that in the control
    throughout the culture period, there was evidence of recovery with
    this compound. The toxicity of the Aroclors was inversely proportional
    to their degree of chlorination. A concurrent investigation of the
    primary productivity of the alga (Hawes et al., 1976b), suggested that
    the productivity of individual cells was stimulated by the Aroclors,
    with a positive relationship between the level of chlorination and the
    effect. However, this did not take into account the differences in
    density between cultures (PCBs reduce growth through cell division)
    and the authors pointed out that the response of the alga to the PCBs
    was not simple. They stated that the density, culture age, and Aroclor
    type were all factors that influenced response.

    Larsson & Tillberg (1975) cultured the green alga  Scenedesmus
     obtusiusculus, in a liquid medium, in concentrations of Aroclor 1242
    ranging between 10 and 1000 g/litre. Growth was reduced at
    concentrations of 300 g Aroclor/litre or more; viability was only
    affected at the highest concentration of 1000 g/litre. Reduced
    phosphate uptake, which was nearly identical in light or in darkness,
    was reduced from 300 g/litre upwards. The authors regarded this as a
    result of an effect on the plasmalemma. At 800 g PCBs/litre, the
    results of some studies suggested an effect of the uncoupling of
    oxidative phosphorylation; at 1000 g/litre, both respiration and
    oxygen evolution were inhibited.

    The effects of various Aroclors on the respiration and photosynthesis
    of the green alga  Chlorella vulgaris were examined by Sinclair et
    al. (1977). Aroclor 1221, at a final concentration in the medium at
    10-4 mol/litre (= 192 mg/litre based on an average relative molecular
    mass for the Aroclor at 192), produced a rate of oxygen production in
    the light of 43% of control levels and a rate of oxygen uptake in the
    dark of 59% of control levels. The Aroclors were dissolved using
    dimethylformamide (DMF) as a solubilising agent; whilst this had some
    effects on the parameters measured, they were very little compared
    with the effects of the Aroclors (94% of control levels). A
    dose-response curve of the effect of Aroclor 1221 on the respiratory
    uptake of oxygen in the dark, in the presence of glucose, showed that
    there was already a marked effect at a concentration of
    10-7 mol/litre and that a maximum (at 50% inhibition) was reached at
    a concentration of 10-6 mol/litre (= 1.92 mg/litre).

    A further series of studies was performed to investigate the
    individual processes of oxygen exchange in the alga. Respiration in
    the dark was investigated in the absence of added glucose, to monitor
    "endogenous" respiration, which was found to be stimulated by Aroclor
    1221 at concentrations of 10-4 mol/litre or more. Net oxygen
    production in the light (photosynthetic oxygen production minus
    respiratory oxygen usage) was reduced at a concentration of
    10-4 mol/litre. At this concentration of Aroclor 1221, there was
    approximately 50% stimulation of endogenous respiration and
    approximately 50% inhibition of net oxygen production. Calculated
    photosynthetic rate was unaffected by the Aroclor. Increasing light
    intensity reduced the effect of Aroclor 1221 on net oxygen production;
    intensities of 8.2  104 ergs/cm2 virtually eliminated the effect.
    Other measures directly or indirectly associated with photosynthesis
    (fluorescence, oxygen evolution in flashing light, the Emerson
    enhancement phenomenon) were not affected by the Aroclor. The authors

    suggested that the photosynthetic apparatus of  Chlorella was
    unaffected by Aroclor 1221; the major, and probably the only, effect
    of the PCBs being a stimulation of endogenous respiration rate.
    Results with Aroclors 1242 and 1268 were consistent with those for
    Aroclor 1221; both inhibited net oxygen production in the light and
    glucose-driven respiration in the dark, but stimulated endogenous
    respiration in the dark (Sinclair et al., 1977).

    Luard (1973) reported inhibition of 14C uptake by the green alga
     Scenedesmus quadricauda at concentrations of Aroclor 1254 as low as
    0.1 g/litre. At this concentration, the Aroclor caused a 20%
    inhibition of 14C uptake, which rose to 65% inhibition at 1 mg/litre.

    A marked effect of the initial number of cells in the incubation tube
    on the toxicity of Aroclor 1254 for the green alga  Chlorella
     pyrenoidosa was demonstrated by Cole & Plapp (1974). At a constant
    concentration of 1 mg Aroclor 1254/litre, the numbers of algal cells
    in the initial incubation medium were as varied as 1, 10, 100, or
    1000 g alga/ml of medium. At the highest inoculation rate, the growth
    of the alga was unaffected by the Aroclor. At lower initial
    inoculation, the rate of growth was reduced to between 11 and 55% of
    control levels. A comparable effect was found using 14C fixation as
    the parameter; with inoculation rates of 1000 g alga/ml medium, the
    Aroclor did not have any effects, whereas lower inoculation rates
    reduced carbon fixation to between 6 and 13% of control levels.

    Mosser et al. (1972) showed that 2 species of freshwater algae
     (Euglena gracilis and  Chlamydomonas reinhardtii) were unaffected
    by 100 g PCBs/litre (type unspecified). Ewald et al. (1976)
    determined the 48-h EC50 on the growth of  Euglena gracilis to be
    4.4 mg/litre for Aroclor 1221 and 55 mg/litre for Aroclor 1232.
    Aroclor 1242 showed no inhibition at 100 mg/litre. Aroclor 1221, at
    the EC50 concentration of 4.4 mg/litre, significantly depressed
    carbon fixation and chlorophyll levels, but did not affect oxygen
    consumption. Uptake of L-leucine was increased 2-fold, but
    incorporation was not affected. Uridine uptake was significantly
    decreased, but thymidine uptake and incorporation were not affected.

    In studies by Glooschenko & Glooschenko (1975), 3 algal species from
    the Great Lakes were cultured with Aroclor at concentrations of 1, 5,
    10, 20, or 50 g/litre. Cell numbers of the diatom  Synedra acus were
    reduced in culture from day 3 of treatment with Aroclor 1242, at 10,
    20, or 50 g/litre, and from day 7, with 1 or 5 g/litre. The green
    alga  Scenedesmus quadricauda showed a lag phase of 3 days in all
    concentrations of Aroclor 1242. After 3 days, exponential growth
    occurred in all treatments, except for the highest dose levels (20 and
    50 g/litre) which showed little or no cell division. A second green
    alga  (Ankistrodesmus falcatus) was more sensitive, showing
    significantly reduced cell numbers at all dose levels of Aroclor 1242.

     Ankistrodesmus was used to examine the relative toxicities of
    different Aroclors. Cell numbers were 57, 66, 36, and 53% of control
    levels for Aroclors 1016, 1221, 1242, and 1248, respectively, after 2
    days of culture. Carbon fixation, estimated as uptake of 14C from
    solution, was 73, 98, 51, and 59% of control levels for the 4
    Aroclors, respectively.

    Dive et al. (1976) cultured the ciliate protozoan  Colpidium campylum
    with purified chlorobiphenyls and with a commercial mixture (Pyralene
    3010). None of the 16 isomers affected the ciliate's growth and
    reproduction at concentrations of 0.01 or 0.1 mg/litre; similarly,
    Pyralene 3010 was not toxic at these concentrations. 2-Mono-
    chlorobiphenyl showed little toxicity for the organism at 1 mg/litre
    and little or no toxicity was demonstrated by the tetrachloro-,
    pentachloro-, and hexachlorobiphenyls at this concentration or at
    10 mg/litre (with the exception of 2,5,2',5'-tetrachlorobiphenyl,
    which inhibited growth considerably, limiting it to about 10% of
    controls at 10 mg/litre). 4,4'-Dichlorobiphenyl was not toxic at any
    concentration tested (up to 10 mg/litre) but both 2,3- and
    2,5-dichlorobiphenyls were toxic, killing all organisms at both 1 and
    10 mg/litre. These results are comparable with other reports that the
    lower chlorinated biphenyls are the most toxic for microorganisms;
    differences in toxicity could not be explained by the differential
    uptake of the different isomers.

    It was reported by French (1976) that the EC50 for growth inhibition
    of Aroclor 1254 on the flagellated protozoan  Crithidia fasciculata
    was 10.5 mg/litre. The PCBs slightly inhibited (after 6 h) and then
    increased (after 24 h) the carbon dioxide evolution of cultures
    utilizing D-glucose. After 6 h exposure, the uptake and incorporation
    of thymidine and uridine (but not L-leucine) were inhibited;
    inhibition was transient and returned to normal after 12 or 24 h. Fine
    structural changes were noted after exposure to PCBs including:
    deterioration of the kinetoplast, mitochondrial or cellular swelling,
    and the presence of concentric membrane arrays. It was concluded that
    cell population growth inhibition was due to disruption of uptake,
    incorporation of nucleic acids, and loss of cell regulatory capacity.

    7.1.2 Marine and estuarine microorganisms

    Bourquin & Cassidy (1975) and Bourquin & Kiefer (1975) investigated
    the effects of Aroclors 1016 and 1242 on 85 bacterial isolates from
    various estuarine environments near Pensacola, Florida. Twenty six of
    the 85 isolates were inhibited to various extents by 0.5 mg of either
    of the PCBs, applied to a disc placed on the surface of an agar plate
    on which the bacteria were growing. Zones of inhibition ranged from 14
    to 20 mm in diameter. Cultures that showed sensitivity to Aroclor 1242
    were also inhibited by Aroclor 1016. Sixty five percent of isolates

    inhibited by 0.5 mg of Aroclor 1242 were still sensitive at 0.1 mg,
    and 58% of those sensitive to Aroclor 1016 at 0.5 mg were still
    sensitive to 0.1 mg. Four isolates were examined in a liquid medium.
    Inhibition of the cultures was characterized by a greatly extended lag
    phase (extended from 2 h to at least 14 h); when growth occurred,
    growth curves were parallel with those of the controls. The
    physiological activity of sensitive and insensitive isolates were
    investigated to try to explain the reasons behind sensitivity. More of
    the sensitive isolates were amylase and gelatinase producers (76 and
    86%, respectively, compared with 33 and 42%, respectively, for the
    whole range of isolates). The significance of this observation is
    unclear. Because of the method of exposure in this screening exercise,
    it is difficult to relate the results to exposure in natural waters
    and to draw conclusions about likely hazards for aquatic bacteria.

    Kleppel & McLaughlin (1980) determined the toxic threshold of Aroclor
    1254 for the estuarine diatom  Skeletonema costatum to be between
    3  10-9 and 3  10-8 g/cell. When the effect of cell density on the
    toxicity of the Aroclor for the organism was examined, maximum
    inhibition occurred with the lowest inoculum rates.

    Michaels et al. (1982) estimated the effects of Aroclor 1254 on
    photosynthesis in the marine diatom  Thalassiosira pseudonana
     ( = Cyclotella nana) by monitoring the uptake of 14C-carbon
    dioxide. The numbers of viable cells in the culture were also
    estimated. Total cell numbers were estimated at regular intervals
    during the 48 h of the experiment and the minimum number of viable
    cells required to produce the increase in numbers between periods
    calculated. This gave an estimate of the viable numbers,
    retrospectively, for each period. Inhibition of 14C-uptake per
    culture, per cell, and per viable cell was evident within 1 h of the
    start of the incubation. By 48 h, the 14C uptake per culture was
    reduced to 0.2% of control levels, per cell, to 13% of control levels,
    and per viable cell, to 34% of control levels. The authors concluded
    that the effect of Aroclor 1254 on the diatom is a combination of
    inhibition of carbon assimilation by individual cells and inhibition
    of cell division.

    In an earlier study, Fisher & Wurster (1973) exposed 2 estuarine
    diatoms  (Thalassiosira pseudonana and  Rhizosolenia setigera) and
    an estuarine green alga  (Dunaliella tertiolecta) to Aroclor 1254 at
    0.1 or 10 g/litre, cultured over 100 h. There was no effect on the
    growth of the alga.  T. pseudonana was unaffected by the PCB at
    0.1 g/litre. The effect of the Aroclor on  T. pseudonana at
    10 g/-litre was dependent on temperature; there was a 17% reduction
    in growth rate at 25C, a 44% reduction in growth rate at 18C, and a
    58% reduction in growth rate at 12C. The growth of control cultures

    was greatest at the highest temperature. The growth of  R. setigera
    was completely stopped by all concentrations of PCB tested, for the
    first 48 h of culture. When growth resumed, the degree of inhibition
    was greater at 10C than at 15C. Fisher et al. (1976) exposed the
    marine diatom  Thalassiosira pseudonana to 1 g Aroclor 1254/litre in
    cultures containing various levels of nitrate nutrient. The toxic
    effects of the PCBs on the growth of the diatom were greatest at low
    nitrogen levels. Analysis of variance showed the diatom to be
    significantly dependent for growth on nitrogen concentration and on
    PCB concentration, and that the PCB effect was significantly nitrogen
    dependent. The authors pointed out that marine phytoplankton are often
    nitrogen-limited in nature and suggested that the effects of
    pollutants, such as PCBs, may, therefore, vary with season. The
    greatest effects are likely to occur during bloom conditions, when
    competition for nutrients is greatest. Fisher (1975) considered that
    the presence of PCBs, even at concentrations far above the maximum
    recorded level in sea water, would not affect the overall carbon
    fixation by phytoplankton. Although the cell division of some
    organisms was adversely affected, enough insensitive species existed
    to compensate for the sensitive species. Species diversity and
    community structure were likely to be affected.

    In a study by Craigie & Hutzinger (1975), 6 marine algae were cultured
    for 6 days in the presence of commercial mixtures of PCBs (Aroclors
    1221 to 1262; Phenoclors DP3 to DP6), PCTs (Aroclor 5460), and
    specific chlorobiphenyls. Each compound or mixture was applied to the
    culture medium at 2 concentrations (1 and 100 mg/litre). The algae
    were representatives of 6 different classes: Bacillariophyceae -
     Skeletonema costatum, Thalassiosira fluviatilis; Chlorophyceae -
     Dunaliella tertiolecta; Chrysophyceae -  Monochrysis lutheri;
    Prainophyceae -  Platymonas sp.; Rhodophyceae -  Porphyridium sp.;
    Xanthophyceae -  Olisthodiscus sp. All were cultured at 20C. Growth
    was estimated by turbidity. The response to particular Aroclors was
    species dependent; the least sensitive species was  Dunaliella and
    the most sensitive was  Olisthodiscus. At 1 mg/litre of the PCB
    mixtures, there was relatively little inhibition of  Dunaliella,
     Platymonas, Skeletonema, or Thalassiosira. Olisthodiscus was
    completely inhibited by Aroclors 1248, 1254, 1260, and Phenoclor DP4.
    The Phenoclor series was more toxic for  Olisthodiscus. than the
    Aroclors.  Porphyridium and  Monochrysis showed intermediate
    sensitivity. Generally, the Aroclors with higher chlorination levels
    were less toxic for all species than the Aroclors with lower
    chlorination levels. Results with pure, specific chlorinated biphenyls
    confirmed that highly chlorinated compounds are less toxic than those
    with only 1-4 chlorine atoms per molecule. Experiments with biphenyls
    containing identical percentages of chlorine showed that biological
    response was highly dependent on the structure of the molecule. The
    2,4,2',4'-tetrachlorobiphenyl was more toxic for both  Dunaliella and
     Olisthodiscus than the 2,5,2',5'- isomer, and both were more toxic
    than either the 2,3,4,5- or the 3,4,3',4'- isomers. The last was not
    toxic for any of the algae, even when added at 100 mg/litre.

    Luard (1973) reported a significant reduction in 14C uptake by the
    estuarine green alga  Dunaliella tertiolecta in the presence of
    Aroclor 1254, at 100 g/litre. The culture showed 14C uptake at 65%
    of control levels. At 1000 g/litre, uptake was further reduced to 59%
    of controls. The 14C uptake was unaffected at 10 g/litre.

    7.1.3  Soil microorganisms

    When Murado et al. (1976) added Aroclors to a liquid or solid medium
    on which the soil microfungus  Aspergillus flavus was cultured,
    mycelial growth was reduced progressively as the dose of Aroclor 1254
    increased from 5 to 50 mg/litre in liquid culture. At 25 mg/litre, the
    dry weight of the mycelium was reduced to 1.4, 3.4, 3.9, 3.3, and
    54.6% of control levels by Aroclors 1232, 1242, 1248, 1254, and 1260,
    respectively. At the same time, the relative RNA content of the
    mycelium increased, rising from a control level of 5.9 g RNA/mg dry
    weight to between 13.2 and 18.6 g RNA/mg dry weight for Aroclors 1232
    to 1254. Aroclor 1260 had no marked effect on RNA. The DNA content was
    not affected by any treatment. Cultures on solid medium showed a delay
    in sporulation and a decrease in the diameter of colonies at doses up
    to 20 g/cm2.

    Glooschenko & Glooschenko (1975) cultured the soil alga  Navicula
     pelliculosa with Aroclors 1016, 1221, 1242, and 1248 at a
    concentration of 20 g/litre. The numbers of cells in the culture,
    after 2 days, were reduced to 66, 53, 46, and 56% of control levels
    for the 4 Aroclors, respectively.

    7.1.4  Plankton communities

    Phytoplankton communities from 2 lakes, one oligotrophic and one
    eutrophic, were exposed to PCBs, as Clophen A50, at 26 g/litre
    (Sdergren & Gelin, 1983). In the community from the eutrophic lake
    with a greater biomass of phytoplankton, measurement of 14C uptake,
    monitored immediately after the addition of the PCBs, showed a
    reduction of 34% compared with the controls. Monitoring carbon
    fixation 16 h later showed the phytoplankton recovering from the
    effect of the PCBs, with only 21% inhibition relative to the controls.
    The results differed in the community from the oligotrophic lake. 14C
    uptake immediately after the addition of the Clophen was 70% less than
    that in the controls; 16 h later, the effect was greater, at 84%
    inhibition. The authors pointed out that these results parallel
    findings in pure culture showing that a high density of organisms
    reduces the effects of PCBs.

    Mosser et al. (1972) first demonstrated the effects of PCBs on both
    single and mixed cultures of marine phytoplankton. Two organisms, a
    marine diatom  (Thalassiosira pseudonana) and a marine green alga
     (Dunaliella tertiolecta) were used. The diatom is sensitive and the
    alga insensitive to PCBs. The application of PCBs (not specified) to
    mixed cultures changed the usual dominance of the diatom into
    dominance of the alga, even at concentrations of the PCBs (1 and
    10 g/litre) that had no discernible effects on the diatom in pure

    The effect of Aroclor 1254 on the relative biomass of 2 species of
    marine diatoms  Phaeodactylum tricornutum and  Cyclotella cryptica
    was examined by Lundy et al. (1984). The diatoms were cultured
    together in either 10 or 20 g PCBs/litre for 6 days. At both
    concentrations of Aroclor, the ratio between the species was shifted
    in favour of  Phaeodactylum. After 6 days, the ratio of
     Phaeodactylum: Cyclotella was 0.8 in the controls and 2.63 in the
    treated cultures (10 g Aroclor 1254/litre).

    Biggs et al. (1979) exposed a mixed culture of 2 marine algae to PCBs
    (Aroclor 1254) at 50 g/litre. In the control culture,  Thalassiosira
     pseudonana became the dominant organism over  Dunaliella
     tertiolecta. After exposure to the PCBs, the  Thalassiosira was
    affected, but the  Dunaliella was not. By day 2 of Aroclor 1254
    exposure,  Dunaliella was the dominant species. Fisher et al. (1974)
    showed a similar effect with the greater effects of PCBs on the
    sensitive diatom  Thalassiosira pseudonana when the organism was in
    competition with other organisms. The effect was also demonstrated
    using natural communities of phytoplankton, when  Thalassiosira was
    also selectively affected.

    Biggs et al. (1978) exposed a natural community of marine
    phytoplankton (from which large detritus and zooplankton had been
    filtered) to Aroclor 1254 at either 5 or 10 g/litre. Cell division,
    chlorophyll-a synthesis, and 14C uptake were monitored, as well as
    particle size. Treatment with the Aroclor reduced community growth
    rates by 20-50%, compared with controls. Growth had not fully
    recovered after 10 days. The 14C uptake was reduced for 6 days. The
    control cultures became dominated by algae larger than 8 m in
    diameter. The PCB-treated cultures showed strong inhibition of these
    larger algal cells and the culture became dominated by small cells.

    In a study by Iseki et al. (1981), a natural community of plankton was
    exposed,  in situ, in a marine environment, to Aroclor 1254 in large
    bags holding 68 m3 of seawater. The Aroclor was added to the bag
    giving an initial concentration in the upper layers of about
    40 g/litre (15 g/litre at 10 m depth). Over time, the levels of PCB
    fell in the upper layer and increased in the lower layer in the bag.

    Six days after addition, the concentrations at all depths were less
    than 15 g/litre. Immediately after addition of the PCBs, the rate of
    sedimentation of the particles increased; these particles were thought
    to be dead or senescent large cells, such as diatoms, and this
    sedimentation was assumed to be responsible for the major part of the
    loss of the PCBs from the water. Zooplankters were eliminated from the
    bags by this level of Aroclor; no recovery was seen over the 21 days
    of the experiment. Large diatoms were selectively eliminated from the
    bags and replaced by small flagellates as the dominant organism.

    Moore & Harriss (1972, 1974) exposed a natural community of plankton
    to PCBs (as Aroclor 1242) at 10 or 25 g/litre. The population was
    collected from natural water, placed in glass bottles suspended
     in situ, and monitored for uptake of 14C, added to the bottles as
    bicarbonate. After incubation, the community was separated into small
    "nannoplankton" and larger "net-plankton" by filtration at a mesh size
    of 53 m. Nannoplankton radiocarbon uptake accounted for 72.6% of the
    total community carbon uptake and was not affected by either of the
    concentrations of Aroclor 1242. The net-plankton uptake of 14C was
    reduced by 56% at 10 g Aroclor/litre and by 58% at 25 g/litre. The
    authors suggested that PCBs at levels found in natural waters would
    alter the species diversity or community structure of microorganisms
    and that this might affect higher levels of the food chain with
    specialist feeders utilizing one type of prey. O'Connors et al. (1978)
    exposed natural communities of phytoplankton,  in situ, in dialysis
    bags in a salt marsh. Large zooplankton herbivores were removed by
    filtration through a mesh. Aroclor 1254 was added to the bags to give
    water concentrations of 1-10 g/litre. Larger diatoms were selectively
    inhibited by the Aroclor, even at the lower dose (1 g/litre). The
    authors suggested that, not only would phytoplankton communities be
    affected by PCBs in natural waters, but that the effect would be
    carried forward through the food chain. Gelatinous predators, such as
    jellyfish, could be selected at the expense of fish, since fish tend
    to depend directly, or indirectly, on the larger phytoplankton.

    Natural communities of phytoplankton from a stream and a reservoir
    were cultured with Aroclors 1232 and 1254 by Kricher et al. (1979).
    Although the algal communities were different in composition, both
    Aroclors exerted similar effects on both communities; primary
    productivity was reduced in a dose-dependent manner. Aroclor 1232 was
    more toxic than Aroclor 1254 at 1 mg/litre. Algal species within the
    populations were differentially affected by the Aroclors; diatoms were
    particularly susceptible and treatment produced disproportionate
    numbers of blue-green algae, such as  Anacystis. The authors pointed
    out that the insensitive species still accumulated PCBs and,
    therefore, formed the basis for the accumulation of the Aroclors in
    aquatic food chains.

    7.1.5  Interactions with other chemicals

    Mosser et al. (1974) investigated the interactions between Aroclor
    1254, DDT, and DDE in a marine diatom  Thalassiosira pseudonana. The
    diatom was cultured for 4 days with either 10 or 50 g Aroclor
    1254/litre, 100 g DDE/litre or 500 g DDT/litre (with each chemical
    alone or in combination). The PCBs alone (at 10 g/litre) and the DDE
    alone had little effect on the growth of the diatom; when combined
    these treatments were synergistic, growth being less than half that of
    the control culture. Higher concentrations of either compound
    increased the inhibitory effect. In contrast, DDT reduced the toxic
    effects of PCBs at higher concentrations (50 g/litre). Treatment with
    the Aroclor alone at 50 g/litre almost stopped the growth of the
    diatom. Simultaneous treatment with DDT at 500 g/litre restored
    growth to 60-70% of control levels. Lower concentrations of DDT had a
    comparable, but reduced, effect. Addition of the DDT to the medium, 12
    or 24 h after culture had begun in the presence of PCBs, also reversed
    the inhibitory effect.

    7.1.6  Tolerance

    Fisher et al. (1973) showed that strains of diatoms, isolated from the
    Sargasso Sea, were more sensitive to the effects of PCBs than isolates
    of the same species, obtained from estuaries or the continental shelf.
    It was suggested by the authors that the difference in sensitivity was
    derived from the variable environment of the estuarine strains; these
    strains are able to cope with wide variations in their living
    conditions, not experienced in the open ocean of the Sargasso, and,
    therefore, were better able to cope with stress from chemical

    When Cosper et al. (1984) compared the sensitivity of clones of 2
    species of diatom  (Asterionella japonica and  Ditylum brightwelii)
    from polluted and unpolluted sites,  Asterionella was less sensitive
    than  Ditylum to the action of PCBs (Aroclor 1254); some strains of
    the former were tolerant of 25 g Aroclor/litre whereas no strains of
    the latter could tolerate this concentration. There was evidence that
    strains of  Asterionella from the polluted site were more tolerant
    than the same species from unpolluted sites. One strain, from the
    polluted site, was tolerant to Aroclor 1254 at 50 g/litre.

    7.2  Toxicity for aquatic organisms

    7.2.1  Aquatic plants

    Mahanty (1975) grew the aquatic angiosperm  Spirodela oligorhiza in a
    sterile culture solution, to which had been added Aroclor 1242, at
    concentrations of 5-100 mg/litre. The numbers of colonies were counted
    throughout the 14-day exposure. The highest dose (100 mg/litre) was

    found to be lethal. At both 25 and 50 mg/litre, though there was some
    growth, the colonies were small and showed morphological differences
    from control colonies including: smaller fronds, in larger numbers
    than usual, and a characteristic striped pattern of chlorosis on the
    fronds. Even at 5 mg/litre, growth was reduced by 50% (as recorded by
    the number of colonies and the fresh weight). Mahanty & McWha (1976),
    using only the 5 mg/litre dose, found reduced growth, an unusual
    striped pattern of chlorosis, and a reduction in the levels of
    chlorophyll and total RNA, but no change in the levels of DNA. When
    Mahanty & Fineran (1976) studied treated (5 mg/litre) and untreated
    fronds of  Spirodela by electron microscopy, they found almost
    complete disorganization of the internal structure of the chloroplasts
    in chlorotic tissue. Organization of other cell components was largely

    7.2.2  Aquatic invertebrates

    The acute toxicity of PCBs for aquatic invertebrates is summarized in
    Tables 27 and 28. Toxicity is very variable between species, even
    closely-related species. For most aquatic invertebrates, there is an
    effect of degree of chlorination of the PCBs, but this is not a direct
    correlation, either negative or positive, the most toxic PCBs often
    being in the mid range of chlorination. Under flow-through conditions,
    the toxicity of PCBs appears to be much higher. Over 96 h, under
    static conditions, LC50 values ranged between 12 g/litre and
    > 10 mg/litre for different organisms and different PCBs.

    MATC (maximum acceptable toxicant concentrations) were set for various
    PCBs by the US EPA (1980). These are expressed as a range between
    no-observed-effect values and lowest concentration tested that
    produced a measurable effect. For  Daphnia magna, these values were
    1.2 and 3.5 g/litre for Aroclor 1248, and 2.5 and 7.5 g/litre for
    Aroclor 1254. For the scud  (Gammarus pseudolimnaeus) values for
    Aroclor 1242 were 2.8 and 8.7 g/litre and values for Aroclor 1248
    were 2.5 and 5.1 g/litre. Aquatic larvae of the midge  (Tanytarsus
     dissimilis) had a no-observed-effect level of 0.5 g/litre and a
    lowest effective concentration at 1.2 g/litre.  Short- and long-term toxicity

    Roberts (1975) exposed the common mussel  (Mytilus edulis) to
    Aroclors 1242 and 1254 and studied byssus formation. The 24- and 48-h
    EC50s for a reduction in the number of mussels byssally-attaching
    were 2.2 and 3.0 mg/litre, respectively, for Aroclor 1254. EC50s
    after exposure to Aroclor 1242 were 0.9 mg/litre for 24 h and
    1.0 mg/litre for 48 h. Duke et al. (1970) maintained oysters

        Table 27.  Acute toxicity of PCB mixtures for freshwater invertebrates

    Organism          Size/age   Stat/   Temperature    Hardness      pH     PCB type    Parameter      Concentration    Reference
                                 flowa      (C)        (mg/litre)b          (Aroclor)                  (mg/litre)

    Scud              mature     flow      15             272         7.4    1242        96 h - LC50    0.01             Mayer & Ellersieck
    (Gammarus                                                                                                            (1986)
                      juvenile   flow      18                                1248        96 h - LC50    0.029            Nebeker & Puglisi
                      juvenile   flow      18                                1242        96 h - LC50    0.073            (1974)

    Scud              mature     stat      21             44          7.1    1248        96 h - LC50    0.052            Mayer & Ellersieck
    (Gammarus         mature     stat      21             44          7.1    1254        96 h - LC50    2.4              (1986)

    Glass shrimp      mature     flow      15             272         7.4    1254        168 h - LC50   0.003            Mayer & Ellersieck
    (Palaemonetes                                                                                                        (1986)

    Crayfish          early      stat      21             44          7.1    1242        168 h - LC50   0.03             Mayer & Ellersieck
    (Orconectes nais) instar                                                                                             (1986)

    Crayfish          early      stat      21             44          7.1    1254        168 h - LC50   0.1              Mayer & Ellersieck
    (Procambarus sp.) instar                                                                                             (1986)
                      immature   stat      12             44          7.5    1254        96 h - LC50    > 0.55

    Table 27.  (cont'd)

    Organism          Size/age   Stat/   Temperature    Hardness      pH     PCB type    Parameter      Concentration    Reference
                                 flowa      (C)        (mg/litre)b          (Aroclor)                  (mg/litre)

    Stonefly          first      stat      10             170         7.2    1016        96 h - LC50    0.61             Mayer & Ellersieck
    (Pteronarcella    year                                                                              (0.42-0.88)      (1986)

    Damselfly         late       flow      15             272         7.4    1242        96 h - LC50    0.4              Mayer & Ellersieck
    (Ischnura         instar     flow      15             272         7.4    1254        96 h - LC50    0.2              (1986)

    Dragonfly         late       stat      21             44          7.1    1242        168 h - LC50   0.8              Mayer & Ellersieck
    (Macromia sp.)    instar     stat      21             44          7.1    1254        168 h - LC50   0.8              (1986)

    a  Stat = static conditions: water not changed during the exposure; flow = flow-through conditions; concentration of
       toxicant continuously maintained.
    b  Hardness expressed as mg CaCO3/litre, unless otherwise stated.

    Table 28.  Acute toxicity of PCB mixtures for marine invertebrates

    Organism         Size/age   Stat/   Temperature   Salinity   PCB type       Parameter      Concentration          Reference
                                flowa      (C)          (%)                                                          (mg/litre)

    Cockle           adult      stat        15                   Aroclor 1248   48 h - LC50   > 10                    Portmann &
    (Cardium         adult      stat        15                   Aroclor 1254   48 h - LC50   > 10                    Wilson (1971)
    edule)           adult      stat        15                   Aroclor 1260   48 h - LC50   > 10
                     adult      stat        15                   Aroclor 1262   48 h - LC50   > 10
                     adult      stat        15                   Clophen A30    48 h - LC50      3.0
                     adult      stat        15                   Clophen A60    48 h - LC50   > 10

    Eastern oyster   adult      flow        28            28     Aroclor 1016   96 h - EC50      0.01                 Mayer (1987)

    Brown shrimp     adult      flow        30            29     Aroclor 1016   96 h - LC50      0.01                 Mayer (1987)

    Brown shrimp     adult      stat        15                   Clophen A60    48 h - EC50   > 10                    Portmann &
    (Crangon         adult      stat        15                   Aroclor 1242   48 h - LC50      1.0                  Wilson (1971)
    crangon)         adult      stat        15                   Aroclor 1248   48 h - LC50      0.3-1.0
                     adult      stat        15                   Aroclor 1254   48 h - LC50      3.0-10.0
                     adult      stat        15                   Aroclor 1260   48 h - LC50   > 10
                     adult      stat        15                   Aroclor 1262   48 h - LC50   > 10
                     adult      stat        15                   Clophen A40    48 h - LC50      0.3-1.0
                     adult      stat        15                   Clophen A30    48 h - LC50      1.0-3.3
                     adult      stat        15                   Clophen A50    48 h - LC50      3.3-10.0

    Table 28.  (cont'd).

    Organism         Size/age   Stat/   Temperature   Salinity   PCB type       Parameter      Concentration          Reference
                                flowa      (C)          (%)                                                          (mg/litre)

    Grass shrimp     1 day      stat        25            25     Aroclor 1016   96 h - LC50      0.15                 Mayer (1987)
    (Palaemonets     3 days     stat        25            25     Aroclor 1016   96 h - LC50      0.021
    pugio)           6 days     stat        25            25     Aroclor 1016   96 h - LC50      0.017
                     9 days     stat        25            25     Aroclor 1016   96 h - LC50      0.019
                     12 days    stat        25            25     Aroclor 1016   96 h - LC50      0.021
                     15 days    stat        25            25     Aroclor 1016   96 h - LC50      0.024
                     18 days    stat        25            25     Aroclor 1016   96 h - LC50      0.037
                     30 days    stat        25            25     Aroclor 1016   96 h - LC50      0.044
                     adult      stat        25            25     Aroclor 1016   96 h - LC50      0.052 (0.046-0.057)
                     adult      flow        30            28     Aroclor 1016   96 h - LC50      0.012
                     1 day      stat        25            25     Aroclor 1242   96 h - LC50      0.015
                     3 days     stat        25            25     Aroclor 1242   96 h - LC50      0.019
                     6 days     stat        25            25     Aroclor 1242   96 h - LC50      0.015
                     9 days     stat        25            25     Aroclor 1242   96 h - LC50      0.017
                     12 days    stat        25            25     Aroclor 1242   96 h - LC50      0.016
                     15 days    stat        25            25     Aroclor 1242   96 h - LC50      0.024
                     18 days    stat        25            25     Aroclor 1242   96 h - LC50      0.034
                     30 days    stat        25            25     Aroclor 1242   96 h - LC50      0.041
                     adult      stat        25            25     Aroclor 1242   96 h - LC50      0.057 (0.048-0.062)

    a  stat = static conditions: water not changed during the exposure; flow = flow-through conditions; concentration
       of toxicant continuously maintained.

     (Crassostrea virginica) at concentrations of 1, 10, and 100 g
    Aroclor 1254/litre and monitored shell growth over a period of 96 h;
    the rates of shell growth were decreased by 19, 41, and 100%,
    respectively. Lowe et al. (1972) found the growth rate (height and wet
    weight) of young oysters  (Crassostrea virginica) to be significantly
    reduced after exposure, in flowing sea water, to 5 g Aroclor
    1254/litre, for 24 weeks. No effects on growth were observed at
    1 g/litre over a period of 30 weeks. Oysters exposed to 5 g/litre
    showed atrophy of the digestive diverticular epithelium and
    degeneration of the vesicular connective tissues of the hepatopancreas
    together with leukocytic infiltration. There was complete tissue
    recovery after 12 weeks in clean water.

    After exposing  Daphnia magna to Aroclor 1254, over a period of 14
    days under static renewal procedures, Maki & Johnson (1975) calculated
    an LC50 of 24 g/litre.

    Nebeker & Puglisi (1974) calculated 3-week LC50 values, under static
    conditions, for a range of Aroclors on  Daphnia magna. Aroclors were
    dissolved in acetone and triton X100 to maintain the PCBs in solution.
    Tests were performed in raw Lake Superior water. Results are presented
    in Table 29. The Aroclors most toxic for  Daphnia had between 48 and
    62% chlorination; the most toxic Aroclor was 1248. Under flow-through
    conditions, renewing the original test concentration of the Aroclor
    continuously, the Aroclors were much more toxic. Two-week LC50 values
    for Aroclors 1248 and 1254 were 2.6 and 1.8 g/litre, respectively,
    while the 3-week LC50 for Aroclor 1254 was 1.3 g/litre. Groups of 40
    scud  (Gammarus pseudolimnaeus) were exposed to various concentrations
    of Aroclor 1242 under flow-through conditions. No animals survived
    exposure for 2 months, at Aroclor concentrations of 26 g/litre or

    Survival at lower exposures was 52% at 8.7 g/litre and 77% at
    2.8 g/litre (control survival was low at 48%).

    Duke et al. (1970) conducted acute, flowing-water bioassays on pink
    shrimp  (Penaeus duorarum). At 0.1 mg/litre, 100% of the shrimps were
    killed within 48 h of exposure to Aroclor 1254. There was no mortality
    after 48 h of exposure to 0.01 mg/litre Aroclor. In long-term
    flowing-water bioassays, Nimmo et al. (1971a) found that Aroclor 1254,
    at a concentration of 0.94 g/litre killed 51% of juvenile pink
    shrimps within 15 days. Juveniles were found to be more sensitive than
    adults; 50% of adults were killed after exposure to 3.5 g/litre over
    35 days. There were no apparent symptoms of poisoning prior to death.

        Table 29.  Toxicity of various Aroclors for Daphnia magna in static testsa

    Aroclor                      3-week             Confidence limits (95%)
                             LC50 (g/litre)

    1221                         180                  (158.0-205.0)
    1232                          72                    (62.6-82.8)
    1242                          67                    (55.4-81.0)
    1248                          25                    (21.4-29.2)
    1254                          31                    (25.8-37.2)
    1260                          36                    (27.7-46.8)
    1262                          43                    (37.0-49.9)
    1268                         253                  (222.0-288.0)

    a  From: Nebeker & Puglisi (1974).

    Striped hermit crabs  (Clibanarius vittatus) were kept in static
    seawater solutions containing 3, 5, 10, 15, 20, 25, or 30 g/litre of
    Aroclor 1254 for 96 h (Stahl, 1979). No deaths were reported, though
    the crabs exposed to the higher concentrations (20, 25, and
    30 g/litre) were less active. Six crabs already exposed to 30 g
    PCBs/litre were then placed in solutions containing 300 g/litre for a
    further 96 h; there were still no deaths.

    Vernberg et al. (1977) exposed fiddler crab  (Uca pugilator) larvae
    to concentrations of 0.1, 1.0, 5, 10, 50, or 500 g Aroclor
    1254/litre, for 96 h. They did not find any effects on survival at 0.1
    or 1.0 g/litre; Aroclor 1254 at 5 g/litre increased deaths by 20%,
    but the increase was not statistically significant. Exposure to 10 g
    PCBs/litre resulted in a 57% reduction in survival of larvae.
    Increasing the PCB concentration to 50 g/litre did not greatly
    increase the effect. The 500 g/litre concentration killed all larvae.
    Increasing exposure time, at 5 g/litre, produced a significant
    reduction in survival after 14 days. Exposure of larvae to Aroclors
    1016 or 1254 at 0.1, 1, or 5 g/litre for periods of up to 168 h, was
    then investigated. No significant effects were found on survival with
    any concentration of Aroclor 1016, for up to 120 h of exposure. After
    168 h, survival rates of larvae exposed to 1 and 5 g/litre were
    reduced to 61 and 66%, respectively. There was no effect at
    0.1 g/litre. Aroclor 1254 did not have any effects on survival at
    concentrations of 0.1 or 1 g/litre. Survival was reduced after

    exposure to 5 g Aroclor 1254/litre for more than 96 h, increasing
    from 60-81% up to 168 h. Exposure to 10 g Aroclor 1254/litre caused
    55% deaths after 120 h; there were no further deaths after 168 h.
    Fifty per cent of adult male crabs, exposed to Aroclor 1254 or Aroclor
    1016 at 50 g/litre, died after 2 days and after 4-6 days,
    respectively. Females (50%) exposed to 50 g/litre survived for 7 days
    after exposure to Aroclor 1016 but for only 4 days after exposure to
    Aroclor 1254.

    Neff & Giam (1977) exposed juvenile horseshoe crabs  (Limulus
    polyphemus) to concentrations of Aroclor 1016 of 10, 20, 40, or
    80 g/litre, for up to 96 days. The crabs were divided into 2 groups:
    group A consisted of juveniles at the late first tailed stage and
    group B, of juveniles at the early second tailed stage. The authors
    calculated LT50s (LT50: median survival time for a given exposure
    concentration) of 20.8 days at 40 g/litre and 20.3 days at
    80 g/litre for group A juveniles. The LT50 for group B crabs at
    80 g/litre was 61 days, but less than 50% had died within 96 days at
    40 g/litre.  Response to temperature and salinity

    In a study by Vernberg et al. (1977), larval fiddler crabs
     (Uca pugilator) were exposed to "sub-lethal" concentrations of
    Aroclor 1254 and 1016 and the conditions of temperature (15-30C) and
    salinity (15-36%) varied. Exposure to 0.01 g Aroclor/litre showed no
    consistent effects of temperature or salinity, though the organisms
    exposed under conditions furthest from the optimum (25C and 30%) were
    more likely to differ significantly from the controls. At the optimum
    temperature and salinity, no deaths were recorded in adult crabs
    exposed to 0.1, 1, 10, or 100 g Aroclor 1254/litre for up to 3 weeks.
    Lowering the salinity or increasing the temperature did not have any
    effects on survival with Aroclor 1254 or 1016 at 5 g/litre. Lowering
    both salinity and temperature (to 5% and 10C) caused 50% of crabs to
    die between 21 and 28 days of exposure to Aroclor 1016 at 5 g/litre,
    but there were no lethal effects of Aroclor 1254 at the same
    concentration. Lowering the temperature further (7C and 5%) reduced
    the 50% survival time to between 5 and 8 days for both Aroclors.

    Nimmo & Barrier (1974) reported some deaths in adult brown shrimp
     (Penaeus aztecus) exposed to a "sub-lethal" concentration of Aroclor
    1254 (3 g/litre), for 7 days, after subjection to salinity shock.
    The shrimp normally adapts readily to the wide range of salinity
    found in its natural estuarine habitat. Roesijadi et al. (1976a)
    exposed the adult grass shrimp  (Palaemonetes pugio) to sub-lethal
    (6.3-8.8 g/litre) and lethal (57.6-76.4 g/litre) concentrations of

    Aroclor 1254 for 96 h, at various salinities. They found little effect
    on haemolymph chloride concentration or osmolarity, chloride space
    (the apparent volume of distribution of chloride ions), or chloride
    exchange kinetics. The shrimp showed an adaptive altered permeability
    to chloride ions at a salinity of 17%, the isotonic point. PCBs did
    not affect this permeability change in adult shrimp. In the juvenile
    grass shrimp, there was a reduction in haemolymph chloride levels at
    low salinities in non-steady state exposures; the PCBs delayed the
    permeability change. This disruption of haemolymph chloride was
    associated with high numbers of deaths, even at the "sub-lethal"
    exposure concentration. It was concluded that juveniles died from
    salinity shock, because of delayed adaptive response. In another study
    (Roesijadi et al., 1976b), grass shrimp were exposed to Aroclor 1254
    at 29.4 g/litre for 96 h at various salinities. No appreciable effect
    was observed on total free amino acid levels in abdominal muscle,
    indicating that intracellular osmoregulation was not a major
    consequence of PCB toxicity, though changes in individual amino acid
    concentrations suggested an altered metabolic state. The authors found
    that blood glycine levels showed large decreases after the transfer of
    the shrimps to clean water, a delayed response to the Aroclor
    exposure.  Reproduction

    Sea urchin  (Arbacia punctulata) eggs were exposed to concentrations
    of Aroclor 1254 of 0.5, 1.0, 5.0, and 10.0 mg/litre (Adams, 1983).
    There was no effect on percentage fertilization, percentage pluteus
    development, or percentage mortality when eggs were exposed at
    fertilization. However, when eggs were exposed 1 h prior to
    fertilization, there was a significant reduction in fertilization
    efficiency at all doses. At all but the lowest dose, there was a
    significant increase in mortality and a significant depression in
    successful pluteus development.

    Maki & Johnson (1975) calculated EC50s for total young produced,
    average brood size, and percentage of days reproducing, during a
    14-day exposure of  Daphnia magna to Aroclor 1254; the results were
    19, 23, and 25 g/litre, respectively.  Daphnia magna were exposed to
    a range of Aroclors, by Nebeker & Puglisi (1974) who estimated
    reproductive impairment, measured as a percentage of surviving young
    relative to controls. The study was conducted over 3 weeks under
    static conditions. Results are presented in Table 30. Reproductive
    impairment matches lethality of the Aroclors (see Table 29); there was
    no indication of reproductive effects of the Aroclors at
    concentrations below those leading to the death of adults or young. No
    young were produced by scud  (Gammarus pseudolimnaeus) exposed to
    8.7 g Aroclor 1242/litre. Scud exposed at 2.8 g/litre produced fewer
    young per surviving adult (4.2), compared with controls (6.8).

        Table 30.  Reproductive impairment of Daphnia magna exposed to Aroclors under
               static conditionsa

    Aroclor         Concentration (g/litre) producing reproductive impairment
                             50%                            16%

    1221                     125                             89
    1232                      66                             53
    1242                      63                             48
    1248                      24                             16
    1254                      28                             18
    1260                      33                             22
    1262                      41                             24
    1268                     206                            162

    a  From: Nebeker & Puglisi (1974).  Moulting

    Several authors have suggested that crustaceans may be more
    susceptible to the toxic effects of PCBs during moult (Duke et al.,
    1970; Wildish, 1970; Nimmo et al., 1971a,b). Fingerman & Fingerman
    (1979) exposed 2 groups of fiddler crabs  (Uca pugilator) at 8 mg
    Aroclor 1242/litre, one group for 40 days, and the other for only the
    first 14 of the 40 days. Eye-stalks were removed on day 15 to increase
    moulting activity. Controls underwent rapid ecdysis, with more than
    50% of the population completing moult within 40 days. Crabs exposed
    to the Aroclor for the full 40 days showed less moulting, with no more
    than 10% of the population completing moult. Crabs exposed for 14 days
    showed 20% of the population successfully moulting. Removal of
    eye-stalks on day 1 of the study produced similar results, in terms of
    numbers of crabs moulting in each treatment, but speeded-up moult in
    the controls. No more Aroclor-exposed crabs moulted after eye-stalk
    removal. In an earlier study, Fingerman & Fingerman (1977) exposed
    fiddler crabs to Aroclor 1242 at 8 mg/litre for 38 days. Either
    eye-stalks or 4 walking legs were removed to stimulate moulting. Both
    control groups underwent ecdysis rapidly. Treated crabs without
    eye-stalks did not undergo any moult. Moulting in those with legs
    removed was much slower than in the controls. The authors also exposed
    crabs to dibenzofuran (1,2,3,4,5,6,7,8-octachlorodibenzofuran) at
    16 ng/litre. This is equivalent to the maximum reported dibenzofuran
    contamination of the Aroclor with the dose equivalent to (in terms of
    the dibenzofuran) the same concentration (8 mg/litre) in the PCB
    mixture. There was only a slight inhibition of moulting caused by the

    Neff & Giam (1977) exposed juvenile horseshoe crabs  (Limulus
     polyphemus) to concentrations of Aroclor 1016 of 10, 20, 40, or
    80 g/litre, for up to 96 days. The crabs were divided into 2 groups:
    group A consisted of juveniles at the late first tailed stage and
    group B of juveniles at the early second tailed stage. ET50s (median
    time for moulting to begin) were calculated between the start of the
    study and the first moult (ET50-1) and between subsequent moults
    (ET50-2 to -n). No moult occurred within 96 days at concentrations of
    Aroclor of 40 or 80 g/litre in either group of crabs. The ET50-1, in
    group A at 10 and 20 g/litre, was not different from controls; the
    ET50-2 was slightly decreased. In group B, 10 and 20 g/litre did not
    affect the ET50-1; 40 and 80 g/litre delayed the onset of the first
    moult by 7 and 9 days, respectively. The ET50-3 were substantially
    decreased at 10, 20, and 40 g Aroclor/litre.  Behaviour

    Hansen et al. (1974a) studied the avoidance response of the pink
    shrimp  (Penaeus duorarum) and the grass shrimp  (Palaemonetes pugio)
    given a choice between clean water and water containing Aroclor 1254,
    at concentrations of between 0.001 and 10 mg/litre. Pink shrimp did
    not avoid any of the concentrations used; grass shrimp only
    significantly avoided the highest dose.  Population structure

    The composition of communities of estuarine animals, in aquaria, were
    studied under different exposures to Aroclor 1254 (0.1, 1.0, and
    10.0 g/litre) for 4 months (Hansen, 1974). The author found that, in
    control groups and the group at the lowest concentration, the
    community was mainly comprised (>75%) of arthropods, mostly the
    amphipod  Corophium volutator. At 1 and 10 g Aroclor 1254/litre, the
    numbers of arthropods decreased and the numbers of chordates
    increased; at 10 g/litre, over 75% of the animals were tunicates. The
    highest concentration of Aroclor decreased the numbers of phyla,
    species, and individuals represented (particularly of amphipods,
    bryozoans, crabs, and molluscs), whereas the numbers of annelids,
    brachypods, coelenterates, echinoderms, and nemertines were
    unaffected.  Interactions with other chemicals

    Maki & Johnson (1975) exposed  Daphnia magna to various combinations
    of DDT (0.2-0.75 g/litre) and Aroclor 1254 (2-24 g/litre). They
    studied adult mortality, total young produced, average brood size, and
    percentage days reproducing, during a 14-day exposure period. The
    effects of one toxicant significantly enhanced the action of the other
    for all test parameters. In the presence of a no-observed-effect level
    of PCB (12 g/litre), the susceptibility of  Daphnia to DDT increased
    by one third. When combined with 0.5 g DDT/litre, the toxicity of
    Aroclor 1254 was doubled.

    In a study by Nimmo & Bahner (1976), the pink shrimp  (Penaeus
     duorarum) was exposed to various combinations of Aroclor 1254
    (0.7-1.1 g/litre), cadmium (640-829 g/litre), and methoxychlor
    (0.9-1.0 g/litre) and numbers of shrimps dying were monitored. The
    results showed no evidence of synergism or potentiation in any

    7.2.3   Fish

    The acute toxicity of PCBs for fish is summarized in Table 31. Values
    for 96-h LC50s vary between 0.008 mg/litre, for the fry of the
    fathead minnow, to > 100 mg/litre for channel catfish. This
    considerable variation is dependent on species and on the PCB mixture
    but appears to depend little on test conditions, such as temperature
    and water hardness. The toxicity of PCBs appears much greater in
    flow-through tests, where the water concentration of the PCBs is
    constantly maintained.

    MATC (maximum acceptable toxicant concentrations) were set by the US
    EPA (1980) and are expressed as a range between the no-observed-effect
    level (based on the results of long-term studies and the sub-lethal as
    well as the lethal effect) and the lowest concentration showing a
    measurable effect. These values, for the fathead minnow, were 5.4 and
    15.0 g/litre for Aroclor 1242; 0.1 and 0.4 g/litre for Aroclor 1248;
    1.8 and 4.6 g/litre for Aroclor 1254, and 1.3 and 4.0 g/litre for
    Aroclor 1260. The early life stage of the estuarine sheepshead minnow
    gave values of 3.4 and 15.0 g/litre for Aroclor 1016 and 0.06 and
    0.16 g/litre for Aroclor 1254.  Short- and long-term toxicity

    The toxicity of Aroclors for 3 species of freshwater fish, over
    exposure of up to 30 days, was systematically investigated by Mayer et
    al. (1977) under flow-through conditions. Results are presented in
    Table 32. Short-term tests consistently underestimate the toxicity of

        Table 31.  Acute toxicity of PCB and PCT mixtures for fish

    Organism/              Size/      Stat/   Temperature  Alkalinityb   Hardnessb   pH    PCB type       Parameter    Concentration
    reference              age        flowa   (C)                                                                     (mg/litre)

    Channel catfish        0.60 g     stat    20                          40         7.4   Aroclor 1016   96-h LC50    > 100
    (Ictalurus             yolk-sac   stat    25                         272         7.4   Aroclor 1016   96-h LC50    0.44 (0.34-0.56)
    punctatus)             2.80 g     flow    17                         272         7.4   Aroclor 1242   96-h LC50    >0.10
    Mayer &                2.80 g     flow    22                         272         7.4   Aroclor 1248   96-h LC50    >0.1
    Ellersieck             2.80 g     flow    22                         272         7.4   Aroclor 1254   96-h LC50    >0.20
    (1986)                 2.80 g     flow    22                         272         7.4   Aroclor 1260   96-h LC50    >0.40

    Atlantic salmon        5.60 g     flow    17                         314         7.6   Aroclor 1016   96-h LC50    0.13 (0.11-0.16)
    (Salmo salar)
    Mayer &
    Ellersieck (1986)

    Brook trout            3.0 g      flow    12                         314         7.6   Aroclor 1016   96-h LC50    >0.80
    Mayer &
    Ellersieck (1986)

    Brown trout            4.60 g     flow    12                         314         7.6   Aroclor 1016   96-h LC50    0.14 (0.11-0.18)
    (Salmo trutta)         1.10 g     stat    13                          44         7.4   Aroclor 1260   96-h LC50    > 24.0
    Mayer &
    Ellersieck (1986)

    Table 31.  (cont'd).

    Organism/              Size/      Stat/   Temperature  Alkalinityb   Hardnessb   pH    PCB type       Parameter    Concentration
    reference              age        flowa   (C)                                                                     (mg/litre)

    Cutthroat trout        2.70 g     stat    9            159           162         7.4   Aroclor 1221   96-h LC50    1.17 (0.96-1.43)
    (Salmo clarki)         2.20 g     stat    9            159           162         7.4   Aroclor 1232   96-h LC50    2.5 (1.72-3.08)
    Mayer &                2.40 g     stat    9            159           162         7.4   Aroclor 1242   96-h LC50    5.42 (3.82-7.68)
    Ellersieck             2.50 g     stat    9            159           162         7.4   Aroclor 1248   96-h LC50    5.75 (5.1-6.5)
    (1986)                 2.50 g     stat    9            159           162         7.4   Aroclor 1254   96-h LC50    42.5 (38.7-46.7)
                           2.70 g     stat    9            159           162         7.4   Aroclor 1260   96-h LC50    60.9 (55.4-67.0)
                           2.40 g     stat    9            159           162         7.4   Aroclor 1262   96-h LC50    > 50
                           2.20 g     stat    9            159           162         7.4   Aroclor 1268   96-h LC50    > 50
                           2.70 g     stat    9                          162         7.4   Aroclor 4465   96-h LC50    > 65
                           2.10 g     stat    9                          162         7.4   Aroclor 5442   96-h LC50    > 50
                           2.90 g     stat    9                          162         7.4   Aroclor 5460   96-h LC50    > 50

    Lake trout             fry        stat    10                         170         7.2   Aroclor 1016   96-h LC50    0.48 (0.39-0.60)
    (Salvelinus            yolk-sac   stat    10                         170         7.2   Aroclor 1016   96-h LC50    0.89 (0.69-1.15)
    Mayer &
    Ellersieck (1986)

    Table 31.  (cont'd).

    Organism/              Size/      Stat/   Temperature  Alkalinityb   Hardnessb   pH    PCB type       Parameter    Concentration
    reference              age        flowa   (C)                                                                     (mg/litre)

    Rainbow trout          0.50 g     stat    12                          44         7.4   Aroclor 1016   96-h LC50    0.14 (0.11-0.16)
    (Salmo gairdneri)      2.50 g     flow    10                         272         7.4   Aroclor 1016   96-h LC50    0.62 (0.42-0.90)
    Mayer &                fry        flow    12                         314         7.6   Aroclor 1016   96-h LC50    0.44 (0.37-0.53)
    Ellersieck             1.80 g     flow    17                         272         7.4   Aroclor 1242   120-h LC50   0.07
    (1986)                 1.80 g     flow    17                         272         7.4   Aroclor 1248   120-h LC50   0.05
                           1.80 g     flow    17                         272         7.4   Aroclor 1254   120-h LC50   0.14
                           1.80 g     flow    17                         272         7.4   Aroclor 1260   96-h LC50    >0.23

    Harlequin fish         10-30 mm   flow                 20             20         8.1   Aroclor 1221   96-h LC50    1.05
    (Rasbora               10-30 mm   flow                 20             20         8.1   Aroclor 1232   96-h LC50    0.32
    heteromorpha)          10-30 mm   flow                 20             20         8.1   Aroclor 1242   96-h LC50    0.37
    Tooby et al.           10-30 mm   flow                 20             20         8.1   Aroclor 1254   96-h LC50    1.1
    (1975)                 10-30 mm   flow                 20             20         8.1   Aroclor 1262   96-h LC50    >100

    Bluegill sunfish       0.90 g     stat    12                          44         7.4   Aroclor 1016   96-h LC50    0.60
    (Lepomis               1.80 g     flow    20                         272         7.4   Aroclor 1016   96-h LC50    0.46 (0.39-0.54)
    macrochirus)           2.20 g     flow    17                         272         7.4   Aroclor 1242   120-h LC50   0.13
    Mayer &                0.80 g     stat    18                          44         7.1   Aroclor 1248   96-h LC50    0.69 (0.48-0.99)
    Ellersieck             2.20 g     flow    22                         272         7.4   Aroclor 1248   120-h LC50   0.14
    (1986)                 0.80 g     stat    18                          44         7.1   Aroclor 1254   96-h LC50    2.74 (1.29-5.81)
                           2.20 g     flow    22                         272         7.4   Aroclor 1254   96-h LC50    0.20
                           2.20 g     flow    22                         272         7.4   Aroclor 1260   96-h LC50    0.40

    Table 31.  (cont'd).

    Organism/              Size/      Stat/   Temperature  Alkalinityb   Hardnessb   pH    PCB type       Parameter    Concentration
    reference              age        flowa   (C)                                                                     (mg/litre)

    Longnose sucker        finger     flow    12                         314         7.5   Aroclor 1016   96-h LC50    0.33 (0.22-0.49)
    Mayer &
    Ellersieck (1986)

    Yellow perch           1.20 g     flow    17                         314         7.6   Aroclor 1242   96-h LC50    >0.15
    (Perca flavescens)     1.10 g     flow    17                         314         7.6   Aroclor 1248   96-h LC50    >0.1
    Mayer &                1.00 g     flow    17                         314         7.6   Aroclor 1254   96-h LC50    >0.15
    Ellersieck             1.20 g     flow    17                         314         7.6   Aroclor 1260   96-h LC50    >0.20

    Fathead minnow         fry        flow    24                                           Aroclor 1254   96-h LC50    0.008
    (Pimephales            fry        flow    24                                           Aroclor 1242   96-h LC50    0.015
    promelas)              3 months   flow    24                                           Aroclor 1242   96-h LC50    0.30
    et al. (1974)

    Cisco (chub)           22 days    flow    7            30-35         40-48             Aroclor 1254   96-h LC50    >10
    (Coregonus sp.)        22 days    flow    7            30-35         40-48             Aroclor 1254   120-h LC50   3.2 (1.9-5.5)
    Passino &
    Kramer (1980)

    Table 31.  (cont'd).

    Organism/              Size/      Stat/   Temperature  Alkalinityb   Hardnessb   pH    PCB type       Parameter    Concentration
    reference              age        flowa   (C)                                                                     (mg/litre)

    Carp (Cyprinus         fry        stat    23-25                                        Kanechlor 300  96-h LC50    1.45
    Kimura et al.

    Guppy (Lebistes        fry        stat    24-25                                        Kanechlor 300  96-h LC50    0.9
    reticulatus)           0.35 g     stat    24-25                                        Kanechlor 300  96-h LC50    3.2
    Kimura et al. (1974)

    a  stat = static conditions: water not changed during the exposure;
       flow = flow-through conditions; concentration of toxicant continuously maintained.
    b  Alkalinity and hardness expressed as mg/litre CaCO3, unless otherwise stated.

        Table 32.  Toxicity of Aroclors for fish (LC50s in g/litre) at 17Ca

    Aroclor        Exposure (days)
                   5      10      15        20        25        30

    Rainbow trout

    1242           67      48      18       10        12         -
    1248           54      38      16        6.4       3.4       -
    1254            -     160      64       39        27         -
    1260            -     326     143       78        49        51

    Bluegill sunfish

    1242            -       -     164      125       120        84
    1248          136     115     111      106       100        78
    1254            -       -     303      260       239       177
    1260            -       -       -        -         -       400

    Channel catfish

    1242            -       -     219      150       132        87
    1248            -     121     121      115       104        75
    1254            -     303     286      293       181       139
    1260            -     535     482      512       465       433

    a  From: Mayer et al. (1977).

    Duke et al. (1970) kept juvenile pinfish  (Lagodon rhomboides)  in
    seawater containing 1, 10, or 100 g Aroclor 1254/litre for up to
    48 h. There were no deaths at any concentration. It was suggested by
    Nimmo et al. (1975) that acute toxicity tests underestimated the true
    sensitivity of marine species; in bioassays lasting 1 week or more,
    Aroclor proved to be 100 times more toxic than acute exposure
    suggested. In tests lasting 2 weeks or longer, Aroclor 1254 was lethal
    for longnose killifish  (Fundulus similis) at 1 g/litre and for
    pinfish and spot  (Leiostomus xanthurus) at 5 g/litre. Hansen et al.
    (1974b) did not find any significant lethal effects in pinfish exposed
    to 100 g Aroclor 1016/litre for 96 h but significant mortality (50%)

    was observed after 33 days at 32 g/litre, and, after 18 days, at
    100 g/litre. Nebeker et al. (1974) exposed the flagfish  (Jordanella
     floridae) to Aroclor 1248 for 40 days. No fish survived at a
    concentration of 18 g/litre and only 35% survived at 5.1 g/litre.
    The fish at these two concentrations almost completely lost their fins
    and tails. Fish survival was not affected at concentrations of
    2.2 g/litre or less. In a study by Defoe et al. (1978), fathead
    minnow  (Pimephales promelas) larvae were exposed, in flow-through
    bioassays, to Aroclors 1248 and 1260 for 30 days; the LC50s were
    calculated to be 4.7 and 3.3 g/litre, respectively.

    Hansen et al. (1976a) fed fingerling channel catfish  (Ictalurus
     punctatus) a diet containing 20 mg Aroclor 1242/kg for 20 weeks. The
    fish showed a reduced weight gain and hypertrophy of the liver. When
    the treated fish were transferred to a control diet for 8 weeks and
    then back to the dosed diet for a further 8 weeks, weight gain and
    liver weights returned to normal levels. No histopathological lesions
    were observed in any of the fish fed PCBs.

    Rainbow trout  (Salmo gairdneri) were fed a diet containing 1, 10, or
    100 mg Aroclor 1254/kg over a period of 330 days (Nestel & Budd,
    1975). No effects on growth rates were seen, but renal lesions were
    observed at all doses; however, they were not dose-related. Foci of
    renal necrosis, with cellular or granular cast formation were seen. A
    significant increase in the number of hepatocytes per unit area in the
    liver was observed at all doses and appeared to be dose-related. A
    reduction of 'white pulp' (lymphatic elements) in the spleen was
    observed at 10 and 100 mg/kg diet. Fish with renal necrosis also had
    reduced splenic white pulp and a reduced white cell count.  Carcinogenicity

    Hendricks et al. (1977) studied the combined effects of Aroclor 1254
    (100 mg/kg diet) and aflatoxin B1 (6 mg/kg diet) in rainbow trout. A
    significantly reduced incidence of liver tumours was observed in the
    combination. There was no retardation of growth in the treated
    animals, but they showed glycogen depletion in hepatocytes,
    hyperaemia, and white pulp depletion in the spleen.

    PCB administration prior to aflatoxin B1 treatment also decreased the
    liver tumour incidence, whereas when Aroclor 1254 was fed after
    exposure of trout embryo to aflatoxin B1, there was no effect on the
    formation of liver tumours (Shelton et al., 1984b).

    When rainbow trout were fed 0, 1, 4, or 8 mg aflatoxin B1/kg diet or
    aflatoxin B1 at these dose levels plus 50 mg Aroclor 1254/kg diet,
    the incidence of hepatic tumours was lower in the groups receiving
    aflatoxin combined with PCBs (Shelton et al., 1984a). The inhibition
    of aflatoxin B1-mediated carcinogenicity was also correlated with the
    decreased bacterial mutagenicity of this compound in the presence of
    an Aroclor 1254-induced drug metabolizing enzyme fraction from fish.
    The potential mechanisms of this interaction were further investigated
    by Shelton et al. (1986) by studying the effects of Aroclor 1254 on
    aflatoxin B1 distribution, metabolism, and adduct formation. The
    results from the  in vivo studies showed that PCB treatment resulted
    in a marked increase in the metabolism of aflatoxin B1 to aflatoxin
    M1 and their glucuronide conjugates. The DNA-adduct levels in the
    PCB-treated fish were 48-96% lower than those in the controls. The
    results in the fish model using aflatoxin B1 as the carcinogen were
    associated with the activity of Aroclor as an inducer of cytochrome
    P-450-dependent monooxygenases (Halverson et al., 1985; Shelton et
    al., 1986).  Effects on developmental stages and reproduction

    Birge et al. (1978) determined LC50s and LC1s for 4 species of fish
    from the fertilization of the eggs to 4 days after hatching. Exposure
    times varied with species dependent on the time taken to hatching of
    the eggs; hatching took 22 days for rainbow trout and 3-4 days for the
    other species. In most species, eggs were considerably less sensitive
    to the toxic effects of Aroclors than larvae. The major exception was
    the rainbow trout for which the duration of exposure of the eggs until
    hatching was considerably longer than for other species. LC50s ranged
    from 0.32 to 11.16 g/litre for the 4 species (up to 4 days after
    hatching) and 4 different PCB mixtures; LC1s ranged from 0.009 to
    0.26 g/litre. Results are presented in Table 33. The rainbow trout
    was the most sensitive species tested.

    Adult minnows  (Phoxinus phoxinus) were fed Clophen A50 at 20, 200,
    or 2000 mg/kg diet (on a dry weight basis), for 40 days (Bengtsson,
    1980). Growth was monitored from day 46 to day 79; a significant
    increase in growth (relative to controls) was seen in fish fed the
    highest dose. Other doses caused increases in growth, but these were
    not significant. Stimulated growth had been observed in previous
    studies with minnows fed Clophen A50 at 0.88-78 mg/kg diet (Bengtsson,
    1979). Between days 127 and 166, the swimming performance of the fish
    was tested using a rotary flow technique. Although the PCBs impaired
    swimming performance, this was not statistically significant.
    Reproduction was monitored from day 235 (first day of spawning) to day
    300. Spawning was delayed in the treated groups by 1 day, 1 week, and
    3 weeks for the 3 treatments (20, 200, and 2000 mg/kg diet),
    respectively. Only the highest dose affected hatchability of eggs,
    which was reduced by approximately 80%.

    Snarski & Puglisi (1976) did not find any adverse effects on survival,
    growth, or reproduction of brook trout  (Salvelinus fontinalis)
    exposed to concentrations of Aroclor 1254 of up to 0.94 g/litre, for
    71 weeks. Survival and growth of alevin-juveniles from exposed parents
    were unaffected for up to 90 days.

    Continuous-flow bioassays were conducted over 8 months on fathead
    minnow  (Pimephales promelas), exposing the fish to either Aroclor
    1242 or 1254 (Nebeker et al., 1974). Reproduction occurred at, and
    below, 5.4 g Aroclor 1242/litre, but spawning and egg production
    were very variable. With Aroclor 1254, reproduction occurred at, and
    below, 1.8 g/litre. Spawning occurred at 1.8 g/litre, but was
    significantly less than spawning at the lower concentrations of 0.23
    and 0.52 g/litre. Egg hatchability and fry survival were good at
    1.8 g/litre. Eggs were more resistant than fry at 15 and 51 g
    Aroclor 1242/litre; with Aroclor 1254 at a concentration of
    15 g/litre, eggs hatched readily, but all fry were dead within 96 h.

    Halter & Johnson (1974) kept coho salmon  (Oncorhynchus kisutch) eggs
    in solutions containing 4.4-56.4 g Aroclor 1254/litre. Exposure
    continued for 2 weeks before, and 4 weeks after hatching, until the
    young were alevins. Hatchability was reduced by 30% at the highest
    concentration of 56.4 g/litre. Survival of alevins was markedly
    higher when eggs were transferred to clean water prior to hatching,
    but there was still 58% mortality in alevins hatched from eggs at the
    highest dose. When the alevins were exposed to the PCBs for 4 weeks
    after hatching, survival was inversely related to exposure
    concentration. No group survived as well as the controls (for example,
    18% died at 4.4 g/litre and 90% at 26 g/litre, or more).

    In a study by Defoe et al. (1978), fathead minnow  (Pimephales
     promelas) were maintained in flow-through bioassays in solutions of
    Aroclors 1248 or 1260, for 240 days (a full life-cycle test).
    Reproduction occurred at all concentrations tested (up to 3 g/litre
    for Aroclor 1248 and up to 2.1 g/litre for Aroclor 1260). The authors
    concluded that PCBs did not produce major effects on reproduction at
    concentrations up to the 30-day LC50 and that reduced populations
    after long-term exposure are largely due to larval mortality.

    Weis & Weis (1982) exposed eggs of the mummichog  (Fundulus
     heteroclitus) to concentrations of Aroclor 1254 of 0.01-10 mg/litre
    (after cleavage had begun). No effect was found on embryonic
    development or hatching and embryonic mortality was negligible.
    Seven-day larval tests also showed no effect on mortality up to the
    highest concentration. However, the authors found approximately 20%
    larval mortality, when larvae were exposed to 5 mg/litre for 72 h,
    after hatching from eggs at the highest exposure concentration. In a

    second group of studies, eggs were exposed to 10 mg Aroclor
    1242/litre; no malformations were found, though there was a consistent
    retardation of hatching. There was a positive correlation between
    hatching rate and female length (length being related to age). Larvae
    exposed to 5 mg/litre showed an average of 45% mortality within 72 h.
    Pre-exposure as eggs to 10 mg/litre greatly increased larval
    mortality; pre-exposure to 1 mg/litre resulted in an intermediate
    response. There was no mortality in control larvae, even when they had
    been pre-exposed as eggs.

    Eggs of brook trout  (Salvelinus fontinalis) were exposed to Aroclor
    1254 (0.043-13 g/litre) for 10 days before hatching, and the fry for
    118 days after hatching (Mauck et al., 1978). Median hatching time,
    egg hatchability, and sac-fry survival were not affected by the PCBs.
    Significantly decreased survival (32% survived) was seen after 48 days
    at 3 g/litre. There was significant mortality of fry after 118 days
    with exposure to concentrations of 3.1 g/litre and above. Growth of
    the trout, as measured by weight, was significantly decreased after 48
    days at concentrations of 1.5 g/litre or more. By the end of the
    study (118 days), no significant differences in weight were seen
    between surviving fry on different treatments. Analysis of the
    backbone composition at this time showed that hydroxyproline and
    phosphorus were significantly decreased by concentrations of
    0.43 g/litre or more and that, at 0.69 g or more/litre, calcium
    levels were significantly increased. Although collagen was
    significantly decreased at 0.69, 3.1, and 6.2 g/litre, it was
    unaffected at 1.5 g/litre; no explanation for the anomaly was

    Schimmel et al. (1974) exposed sheepshead minnow  (Cyprinodon
     variegatus) eggs, immediately after fertilization, to Aroclor 1254
    at concentrations of between 0.1 and 10 g/litre. The fertility of
    eggs was unaffected. Hatching was significantly reduced (by 30%) only
    at the highest exposure. Survival of fry to 2 weeks was significantly
    reduced at concentrations of 0.32 g/litre or above (30% survival at
    0.32 g/litre and 9% survival at 10 g/litre). The 3-week LC50 for
    embryo-fry was calculated to be 0.93 g/litre for Aroclor 1254. Many
    of the dying fish exhibited fin rot. Exposure of juveniles and adults
    to the same concentrations of the Aroclor produced 24% mortality in
    juveniles at the highest exposure rate and no mortality in adults.
    Some fin rot was seen in adults.

        Table 33.  Toxicity of PCBs for the embryolarval stages of fisha,b

    Organism                Aroclor 1016                      Aroclor 1242                      Aroclor 1254
                            LC1             LC50              LC1              LC50             LC1              LC50
                            (g/litre)      (g/litre)        (g/litre)       (g/litre)       (g/litre)       (g/litre)

    Channel catfish         0.08            11.16             0.14             4.24             0.05             1.76
    (Ictalurus punctatus)                   (9.93-12.97)      (0.07-0.23)      (3.32-5.34)      (0.02-0.09)      (1.36-2.24)

    Goldfish                0.10            13.21             0.04             2.64             0.02             1.18
    (Carassius auratus)                     (10.63-16.43)     (0.01-0.08)      (1.89-3.61)      (0.01-0.04)      (0.84-1.61)

    Rainbow trout           0.011           1.08              0.01             1.03             0.009            0.32
    (Salmo gairdneri)       (0.003-0.027)   (0.7-1.56)        (0.002-0.025)    (0.67-1.51)      (0.003-0.02)     (0.22-0.45)

    Redear sunfish          0.26            7.82              0.19             3.56             0.02             0.53
    (Lepomis microlophus)   (0.1-0.51)      (5.74-10.35)      (0.08-0.35)      (0.65-4.66)      (0.01-0.04)      (0.39-0.7)

    a  From: Birge et al. (1978).
    b  Exposure under static conditions, but with water renewed every 12 h. Exposure was initiated 2-6 h after spawning (except
       for rainbow trout where exposure was initiated 15 min after fertilization) and continued to 4 days post-hatching. Hatching
       times varied: 22 days for rainbow trout (13.5-14.3C); 3 days for catfish (29-31C); 3-4 days for goldfish and sunfish
       (20-24C); hardness 90-115 mg/litre; pH 7.6-8.1.

    In another study, Hansen et al. (1975) exposed embryos, fry,
    juveniles, and adults of sheepshead minnow  (Cyprinodon variegatus)
    to concentrations of Aroclor 1016 of 0.1-10 g/litre, for 28 days, in
    intermittent flow bioassays. No effects on survival were observed
    during this period. When exposed to concentrations of 32 or
    100 g/litre, there was high mortality in eggs, juveniles, and adults;
    all were killed at 100 g/litre. The authors calculated that the
    28-day LC50s for juveniles and adults were 20 and 19 g/litre,

    Freeman et al. (1982) fed Atlantic cod  (Gadus morrhua) diets
    containing 1-50 mg Aroclor 1254/kg for 5.5 months. Altered steroid
    biosynthetic patterns  in vitro were observed in the testes and head
    kidneys (adrenal equivalent) of dosed fish. Histological examination
    revealed abnormalities in the testes, gills, and livers. Testicular
    abnormalities included derangement of lobules, hyperplasia of lobule
    walls, and disintegration and/or fatty necrosis of spermatogenic
    elements. In fish fed at dietary rates of 5-50 mg/kg, hyperplasia of
    the epithelial layer of the secondary lamellae of the gills was noted.
    Fatty degeneration of the liver was observed in all treated fish.
    Similar testicular abnormalities were observed by Sangalang et al.
    (1981), but only in sexually mature individuals or at a stage of rapid
    spermatogenic proliferation.  Physiological and biochemical effects

    Coho salmon  (Oncorhynchus kisutch) were fed diets containing a
    mixture of PCBs (1:4, Aroclor 1242:1254) at a concentration of 50 or
    500 mg/kg dry feed (Leatherland & Sonstegard, 1978). Serum
    triiodothyronine (T3) levels were significantly reduced after 3
    months, in fish fed the highest dose. Thyroxine (T4) levels were not
    affected. After 3 months, the T3:T4 ratio was significantly higher in
    fish fed the highest dose than in control and low-dose fish. Fish on
    500 mg/kg had significantly lower body weights than controls by the
    end of the study. A mixture of 50 mg PCBs/kg and 5 mg mirex/kg
    significantly reduced serum triiodothyronine and thyroxine levels over
    a period of 3 months, but did not affect the T3:T4 ratio. Leatherland
    & Sonstegard (1979) fed rainbow trout  (Salmo gairdneri) diets
    containing Aroclor 1254 at 500 mg/kg dry feed, for up to 2 months. The
    PCBs did not have any significant effects on thyroid histology or on
    serum thyroid hormone levels. Liver weights, total liver lipid
    content, and carcase lipid content were significantly greater in
    treated fish. Mayer et al. (1977) fed fingerling coho salmon
     (Oncorhynchus kisutch) 1.45-14 500 g Aroclor 1254/kg body weight
    per day, for 260 days. Channel catfish  (Ictalurus punctatus) were
    fed the same Aroclor at rates of 48 and 480 g/kg body weight per day
    for 193 days. Thyroid activity (as measured by 125I uptake) was

    significantly stimulated at dose rates of 14.5 g/kg per day, and
    above, in coho salmon. Stimulation ranged from 52%, at 14.5 g/kg per
    day, to 119%, at 14 500 g/kg per day, compared with controls. In
    catfish, both dose rates of Aroclor 1254 caused significant increases
    in thyroid activity, whereas other Aroclors (1232, 1248, and 1260) did
    not have any significant effects on the thyroid at the same dose
    rates. Folmar et al. (1982) injected yearling coho salmon
     (Oncorhynchus kisutch) intraperitoneally with a total of 150 g
    Aroclor 1254/kg body weight (2 injections, 10 days apart), prior to
    smoltification. Over a 6-week period, the authors did not find any
    significant effects on gill Na-K ATPase activity. However, the PCBs
    did alter the normal developmental patterns of thyroxine; there was a
    delay in the normal increase in circulating thyroxine levels.
    Triiodothyronine levels were significantly elevated after
    approximately 3 weeks, but then fell to well below control levels
    after 6 weeks. Fish were transferred to seawater; there was no
    significant effect on the gill Na-K ATPase activity in the treated
    group, but there was a significant increase in mortality (6%). Ten per
    cent of fish, dosed with PCBs and placed in sea water containing No 2
    fuel oil at 700 g/litre, died; this was an additive effect.

    Fingerman (1980) kept Gulf killifish  (Fundulus grandis) in a
    seawater solution containing Aroclor 1242, 1254, or 1268 at
    8 mg/litre, for up to 28 days. The author removed the lower half of
    the caudal fin to study fin regeneration. No significant effects were
    observed with either Aroclor 1242 or 1254. With Aroclor 1268, a
    significant decrease in regeneration rate was observed after 28 days,
    when the study was conducted in the spring, and after 7 days, in the
    autumn. No differences were found at other sampling times.

    Rainbow trout  (Salmo gairdneri) were administered capsules
    containing 0.173 g Aroclor 1254 every second day over a 6-day period
    (Kiessling et al., 1983). Two to 4 weeks after the last capsule,
    isolated gills were perfused. There was no significant difference in
    adrenergic response in the gill vascular bed and no significant
    difference in the "oxygen transfer factor" (% changes in oxygen in
    saline from dorsal aorta before, and after, addition of adrenaline).
    Similarly, there was no effect on muscle glycogen content.

    Johansson et al. (1972) dosed brown trout  (Salmo trutta) twice, 4
    days apart, with 5 mg Clophen A50/kg body weight, either by capsule or
    by intramuscular injection. The fish were fed for 43 days, starved for
    116 days, and fed again for another 87 days. Metabolic analysis of the
    fish was undertaken on day 43 and at the end of the study. A
    significant increase in body weight was noted at the end of the study,

    but not after 43 days. Blood glucose, muscle glycogen, and the
    liver-somatic index (liver weight as a ratio of body weight), which
    had all increased significantly after 43 days, decreased significantly
    by the end of the study (compared with controls). Both haematocrit and
    haemoglobin levels had significantly decreased after 43 days but, by
    the end of the study, were not significantly different from those of
    the controls.

    In a study by Camp et al. (1974), fingerling catfish  (Ictalurus
     punctatus) were kept in water containing Aroclor 1254 at 8 mg/litre.
    There was a significant increase in the serum transaminase activity of
    the fish after 4 h. The cortisol content of the serum was depressed,
    but not significantly so. The sodium:potassium ratio was constant.

    Merkins & Kinter (1971) exposed the killifish  (Fundulus heteroclitus)
    to concentrations of Aroclor 1221 of 7.5, 25, or 75 mg/litre. No fish
    died, at the lowest concentration, over a period of 4 days; 50% died
    at 25 mg/litre within 24 h and 88% at 75 mg/litre, within the same
    period. Serum osmolality, and serum ion levels (Na and K) were then
    measured. Within 6 h at 75 mg/litre, blood osmolality significantly
    increased, but sodium and potassium ions in the blood were not
    affected. After 24 h at 75 mg/litre, there was also a significant
    increase in sodium ions, but no effects on potassium ions. None of
    these parameters were affected by exposure at 25 mg/litre for 24 h.  Behavioural effects

    Fingerman & Russell (1980) exposed male Gulf killifish  (Fundulus
     grandis) to Aroclor 1242 at 4 mg/litre for 24 h. A significant
    reduction in whole-brain levels of both noradrenalin and dopamine were
    reported over this period. The swimming activity of the fish was
    monitored by counting the number of times they crossed lines marked on
    the bottom of the tank within a 10-min period, after exposure to the
    Aroclor for 24 h. Activity was significantly increased and remained so
    for a further 2 days.

    The avoidance response was studied by Hansen et al. (1974a) in the
    sheepshead minnow  (Cyprinodon variegatus), the pinfish  (Lagodon
     rhomboides), and the mosquito fish  (Gambusia affinis), when given
    a choice between clean water and water containing Aroclor 1254 at
    0.001, 0.01, 0.1, 1.0, or 10 g/litre. Sheepshead minnow did not avoid
    any concentration; pinfish avoided only the highest concentration of
    the Aroclor. Mosquitofish significantly avoided concentrations of 0.1,
    1.0, and 10 g/litre.

    Peterson (1973) did not find any effect on temperature selection in
    Atlantic salmon  (Salmo salar) exposed to 2 mg Aroclor 1254/litre,
    for 24 h prior to a horizontal temperature gradient test. Similarly,
    Miller & Ogilvie (1975) did not find any effect of Aroclor 1254 on
    temperature selection when brook trout  (Salvelinus fontinalis) were
    exposed to a water concentration of 25-100 mg/litre, for 24 h, prior
    to temperature gradient tests. Even at a concentration of 100 mg/litre
    for 48 h, which was sufficient to cause some mortality, temperature
    selection was still unaffected.  Interactions with other chemicals

    Halter & Johnson (1974) found that the median survival time for coho
    salmon  (Oncorhynchus kisutch) fry, exposed to Aroclor 1254 at
    32.2 g/litre, was greater than 336 h. When fry were exposed to
    mixtures of Aroclor and DDT, for 2 weeks, the survival times were
    always similar to the more rapid reaction time found for DDT alone.
    The authors suggested that this indicated the lack of an additive

    7.2.4  Amphibians

    Tadpoles  (Rana chensinenis) were maintained in water containing PCBs
    (as Kanechlor 300) at 0.5, 5.0, 50, or 500 g/litre. At the 2 highest
    doses, all individuals died rapidly. The time of onset of lethality
    was related to dose level; at 5 g/litre, death occurred between 15
    and 21 days, and, at 0.5 g/litre, on the thirty-second day after
    first exposure. Growth at 0.5 and 5.0 g/litre did not differ from
    that of controls. Tail abnormalities were found, but there was no
    correlation with PCB concentrations and a NOEL could not be
    established. The mechanism by which PCBs caused tail malformation was
    not known (Hasegawa, 1973).

    Birge et al. (1978) conducted embryo-larval bioassays on 3 species of
    amphibia. Exposure to various PCBs was maintained from 2-6 h after
    spawning to 4 days after hatching, using static renewal procedures
    (Table 34). Toxicity increased with increasing chlorination, the
    leopard frog being the most sensitive species with an LC50 of
    1.03 g/litre after exposure to Aroclor 1254. The authors also
    calculated an LC1 value from the same study; the leopard frog and
    American toad were equally sensitive at 0.02 g/litre. The eggs were
    much less sensitive to the PCBs than the hatched larvae, with LC50s
    ranging from 3.5 to 250 g/litre, for 3 different Aroclors and
    Capacitor 21, and 3 species of tadpole.

        Table 34.  Toxicity of PCBs for the tadpoles of amphibiansa,b

    Organism           Aroclor 1016                  Aroclor 1242                  Aroclor 1254                  Capacitor 21

                       LC1           LC50            LC1            LC50           LC1            LC50           LC1            LC50
                       (g/litre)    (g/litre)      (g/litre)     (g/litre)     (g/litre)     (g/litre)     (g/litre)     (g/litre)

    American toad      0.35          7.16            0.03           2.71           0.02           2.02           0.21           9.97
    (Bufo americanus)  (0.15-0.64)   (5.39-9.34)     (0.01-0.06)    (1.91-3.75)    (0.01-0.05)    (1.44-2.77)    (0.08-0.42)    (7.21-13.53)

    Fowler's toad      0.18          27.72           0.22           12.09          0.07           3.74           0.55           28.02
    (Bufo fowleri)     (0.09-0.33)   (21.77-35.08)   (0.12-0.36)    (9.74-14.91)   (0.04-0.11)    (2.98-4.64)    (0.3-0.91)     (22.59-34.47)

    Leopard frog       0.1           6.19            0.04           2.13           0.02           1.03           0.03           2.87
    (Rana pipiens)     (0.05-0.16)   (4.95-7.69)     (0.02-0.06)    (1.72-2.63)    (0.01-0.03)    (0.83-1.27)    (0.02-0.06)    (2.29-3.57)

    a  From: Birge et al. (1978).
    b  Exposure under static conditions, but with water renewed every 12 h. Exposure was initiated 2-6 h after spawning and continued to
       4 days post-hatching. Hatching times varied between 3 and 4 days, therefore, exposure varied between 7 and 8 days.
       Temperature 20-24C; hardness 90-115 mg/litre; pH 7.6-8.1.

    7.2.5  Aquatic mammals

    Following-up on field reports of the reproductive effects of PCBs on
    seal reproduction (see section 7.4.4), Reijnders (1986) conducted a
    study on captive common seals  (Phoca vitulina) fed fish contaminated
    with PCBs. The contaminated diet produced an average daily intake of
    PCBs of 1.5 mg compared with the control level of 0.22 mg/day. Twelve
    female seals were used as controls and 12 as the treated group. Blood
    samples were taken regularly and assayed for circulating steroid
    hormones progesterone and estradiol. Females were mated with undosed
    males. Of the 12 females in the control group, 10 became pregnant; all
    12 ovulated. Only 4 females became pregnant out of the 12 fed the
    PCB-contaminated diet; again all 12 ovulated. Throughout the breeding
    cycle, no significant differences were found between the hormonal
    profiles of pregnant animals in the treated and control groups. No
    significant differences were observed in progesterone levels in the
    treated and control groups, despite the fact that many fewer treated
    females became pregnant. However, a rise in estradiol levels in
    non-pregnant females in the control group was not found in
    non-pregnant females in the treated group, suggesting a difference in
    non-pregnancy in treated females. The effects of PCBs occurred only
    late in the breeding cycle at the time of implantation of the embryo.
    Seals, like mink, show delayed implantation of the embryo as a normal
    component of the annual reproductive cycle. No conclusions about the
    mechanism of action could be drawn.

    Brouwer et al. (1989) fed common seals  (Phoca vitulina) on a diet of
    polychlorinated biphenyl-contaminated fish (average daily intake
    1.5 mg PCBs) for almost 2 years. Significant reductions in levels of
    plasma total and free thyroxine, triiodothyronin, and retinol were
    found compared with those in seals maintained on a "low" contaminated
    diet (average daily intake 0.22 mg PCBs). It should be noted that the
    diet consisted of fish contaminated in the environment and not dosed.
    No attempt was made to analyse levels of retinol in the different fish
    diets. The "high" contamination group were caught in the Wadden sea
    and the "low" contamination group in the north-east Atlantic. When the
    fish were analysed for other likely contaminants, it was found that
     pp'-DDE also showed higher levels in the "high" contamination group
    than the "low" contamination group; average daily intakes of  pp'-DDE
    were estimated to be 0.4 mg and 0.13 mg, respectively.

    7.3  Toxicity for terrestrial organisms

    7.3.1  Plants

    Aroclor 1254 was applied to soil at rates of 10, 100, or 1000 mg/kg
    (Weber & Mrozek, 1979). Both soybean  (Glycine max) and rescue
     (Fescue arundinacea) were grown in the soil, from seed, for up to 26
    and 42 days, respectively. The height of the soybean plants and the
    fresh top weights of both plants were measured. PCBs applied to the

    soil significantly reduced height and fresh top weight of soybean
    plants, only at the highest rate of application. Low rates were
    inhibitory, but not significantly so. Aroclor applied to the soil also
    reduced the fresh top weight of rescue at the highest rate of
    application (1000 mg/kg); lower application rates of the Aroclor did
    not have any effects. The addition of activated carbon to the soil
    (3.7 tonnes/ha; approximately 3333 mg/kg) annulled the inhibitory
    effect of PCBs. The Aroclor also inhibited the uptake of water by the
    soybean in proportion to the dose applied to the soil; water uptake
    was monitored between 21 and 25 days after sowing of the seed. The
    reduction in water uptake over this 5-day period was 12%, with the
    application of 1 mg Aroclor 1254/kg soil, rising to 52% at 1000 mg/kg
    soil. Again, the effect on water uptake was eliminated by the addition
    of carbon to the soil (1% rising to 4% inhibition over the dose

    Continuation of the experiment through a second and third crop of
    soybeans on the same soil, without further addition of Aroclor 1254
    (Strek et al., 1981), showed similar effects on the height, top fresh
    weight, and water uptake of the plants, reflecting the persistence of
    the Aroclor. However, there were no significant effects at doses lower
    than 1000 mg/kg, with the exception of reduced height in the third
    crop, seen at all dose levels. All effects were eliminated by the
    addition of activated carbon to the soil. The same authors found that
    beet  (Beta vulgaris) was significantly affected by 1000 mg Aroclor
    1254/kg, using the same parameters of water uptake, height, and fresh
    top weight between 14 and 56 days after sowing. Doses of 100 mg/kg or
    less did not have any significant effects. Effects at the highest dose
    were again eliminated by the addition of activated carbon to the soil.
    Growth parameters, taken at harvest, showed no apparent inhibition of
    corn  (Zea mays) or sorghum  (Sorghum bicolor) by Aroclor 1254 over
    the same dose range. There was, however, a reduction of plant height
    over the first 5 days of growth at 100 and 1000 mg/kg in corn, but the
    plants recovered.

    Mrozek et al. (1983) grew  Spartina alterniflora plants in mud or
    sandy soils in the presence of 2.2 g PCBs/kg (54% chlorine similar to
    Aroclor 1254), admixed with the soil, over a 6-week period. Plants
    grown in sand showed significantly reduced values for cumulative
    change in height (approx. 30%) and the number of live leaves per stem
    (approx. 25%), and increased values for the number of stems per plant
    (approx. 300%), whereas plants grown in mud showed a significantly
    reduced value for cumulative change in the number of stems per plant
    (approx. 75%). Mud-grown plants also exhibited an altered biomass
    distribution, as indicated by the aerial:below ground biomass ratio,
    which increased from 1.2 to 1.5, on a dry weight basis.

    7.3.2  Terrestrial invertebrates

    Hatch & Allen (1979) observed the behaviour of the snail  (Cepeae
     (=Helix) nemoralis), with regard to the rasping of conspecifics'
    shells to obtain calcium. On a low calcium diet (0.53 mg calcium/kg),
    snails showed an increased tendency to rasp the shells of other
    snails, in order to obtain calcium. The best indicator of this
    behaviour was found to be the counting of holes bored completely
    through the shell. This behaviour was not seen with a high calcium
    diet (250 mg calcium/kg). The addition of PCBs, as a mixture of
    Aroclors 1016 and 1254, to the high calcium diet at a rate of 0.5,
    1.0, or 5.0 mg/kg increased the number of snail shells penetrated by
    other snails. Penetration increased in a dose-dependent manner with 2,
    5, and 7% penetration for the 3 dose rates, respectively. Damage to
    shells, without actual penetration, also increased with PCB treatment
    from the low level found on the control diet to between 16 and 21% on
    the PCB diet. No clear dose-dependent effect was seen using this
    method of assessing damage. Fourth instar nymphs of the grasshopper
     (Chorthippus brunneus), were dosed topically with the PCB mixture
    Aroclor 1254 (Moriarty, 1969). A single dose of either 12.5, 50, or
    200 g/insect was applied in a volume of 1 l of 1,4-dioxan. No
    sublethal effects were detected on either development or reproductive
    potential. At the highest dose, there appeared to be a latent toxicity
    that could be correlated with the mobilization of lipids at moult;
    more moulted males died than unmoulted males over the test period.
    Females took longer to moult than males and showed a distinctly
    bimodal distribution in time to death, with a similar correlation
    between toxicity and moult. Males showed 46% mortality and females,
    41%, after treatment with 200 g Aroclor 1254. Fungal infection
    affected insects on the lower doses and mortality figures are,
    therefore, unreliable.

    Lichtenstein et al. (1969) exposed  Drosophila melanogaster to the
    dry residue of various PCBs. They exposed flies to Aroclors 1221,
    1232, 1242, and 1248 at 200 or 800 g. No mortality was observed after
    48 h at 200 g. At 800 g, there was an increase in mortality with
    decreasing chlorination (after a 48-h exposure to Aroclor 1221, 92%
    had died; only 45% died after exposure to Aroclor 1248 over the same
    length of time). No mortality was observed after a 48-h exposure to
    2000 g of Aroclors 1254, 1260, 1262, or 1268. In a separate study,
    the authors treated houseflies  (Musca domestica) topically with
    either 10 or 20 g (in 2 l of acetone); mortality was assessed after
    24 h. Results were comparable with those from the study on
     Drosophila; deaths were dose related and occurred with Aroclors up
    to 1254, where mortality was 10% at the higher dose. Aroclors of
    higher chlorination than 1254 had no effect. Lower chlorinated
    Aroclors had the greatest effect with more deaths at the highest dose

    (20 g) and lowest chlorination (Aroclor 1221) than with any other
    treatment (43% killed). Plapp (1972) found that the 24-h LC50 for
    Aroclor 1254 in the housefly  (Musca domestica) was > 3000 g/jar
    where the Aroclor was added to a container in acetone, which was dried
    before the addition of the flies. This was true for both
    DDT-susceptible and DDT-tolerant strains. The same author reported a
    powerful synergistic effect between carbaryl and Aroclor 1254; the
    LC50 with carbaryl alone was calculated to be 1386 g/jar and that
    for carbaryl:PCB in the ratio of 1:5, 96 g/jar. The Aroclor was as
    powerful a synergizing agent as piperonyl butoxide.

    Youssef et al. (1974) hatched eggs of the housefly  (Musca domestica)
    on a medium of paper tissue dosed with 0.808 g of Aroclor 1254 per
    200 g of tissue. The adult flies hatching from the eggs were examined
    using the electron microscope for effects on the male reproductive
    tissue. The PCBs induced nuclear and flagellar abnormalities in
    developing spermatids. Spermatid nuclei failed to elongate and
    membranes originating from the nuclear envelope formed invaginations
    into the nucleus. These resulted in the appearance of cytoplasmic
    inclusions in the nucleus. Spermatid flagellae contained an abnormal
    number of axonemes and mitochondrial derivatives; abnormal spermatids
    did not coil and degenerated.

    In a study by Wasilewska et al. (1975), female nematodes
     (Acrobeloides nanus) were exposed to Aroclor 1254. Initially, 60 g
    of the Aroclor were added to a petri dish (on the surface of agar) in
    which there were 20 nematode worms. The nematodes fed on a culture of
    bacteria introduced to the agar at the same time as the worms. After 5
    days of exposure, eggs and adult nematodes were counted and adult
    weights determined. No significant effects were found. In a second
    study, over a longer period (10 days), nematodes were exposed to 15,
    30, or 60 g of the Aroclor. Adverse effects increased with dose, and
    even at the lowest dose, the number of adults was reduced from 123 to
    76, the number of eggs from 539 to 288, and the weight of adults from
    18.9 to 9.4 g. At the highest dose of 60 g per dish, the number of
    adults was 32, the number of eggs, 37, and the weight of adults,
    4.9 g.

    7.3.3  Birds

    Five-day dietary LC50s for PCBs in birds ranged from 604 to >
    6000 mg/kg diet (Table 35). Generally, the oral single dose LD50 and
    the dietary LC50 data are similar to those for mammals. PCBs are less
    toxic for birds than other organochlorines, such as DDT and its
    metabolites and the chlorinated cyclodienes.

    The toxicity of Aroclors in birds increases with the percentage
    chlorination (generally reflected in the final 2 digits of the Aroclor
    number), according to the data of Hill et al. (1975) and Hill &
    Camardese (1986). Hill et al. (1974) noted that the toxicity of
    Aroclors is not simply a reflection of the chlorine content of the
    different Aroclors. They adjusted the dietary content of the Aroclors
    to a constant dietary chlorine level and found the same increased
    toxicity with higher Aroclor numbers.

    Dahlgren et al. (1972) reported some mortality in sub-adult pheasants
    after regular oral doses of Aroclor 1254 ranging from 10 to 210 mg.
    Mortality was related to both dose and body weight; heavier birds
    lived longer, though they lost a greater proportion of body weight. A
    sudden heavy intake of PCBs led to high brain residues. Brain residues
    were best correlated with death; residues in the brain of about
    300-400 mg/kg were considered by the authors to be diagnostic of acute
    poisoning and death (Dahlgren et al., 1972). Stickel et al. (1984)
    concluded that similar levels in the brain killed red-winged
    blackbirds, starlings, brown-headed cowbirds, and grackles. Intake of
    lower doses of PCBs over long periods does not lead to such high brain
    residues; the cause of death after long-term exposure appears to be
    oedema and related symptoms.  Short-term toxicity

    Hurst et al. (1973) observed differential toxicity of PCBs between
    bobwhite quail hens and cocks. This differential was eliminated when
    the tests were conducted on birds not in the breeding condition and
    with short daylengths. Females survived better than males, only when
    they were laying eggs, and survival was well correlated with the
    numbers of eggs produced. The authors concluded that females reduce
    their exposure to PCBs by eliminating the compound in the eggs.

    When Koeman et al. (1969) fed Japanese quail a diet containing 2000 mg
    Phenochlor DP6/kg, all the dosed birds died between 6 and 55 days of
    dosing. The quail developed hydropericardia at this dose level. Vos &
    Koeman (1970) fed one-day-old cockerels a diet containing PCBs at
    400 mg/kg, for 60 days; the PCBs were in one of the following forms,
    Phenochlor DP6, Clophen A60, or Aroclor 1260 (all 3 are 60%
    chlorinated). The mean survival time was calculated to be 24.3 days
    for Phenochlor and 20.5 days for Clophen, only 3 out of 20 birds died
    on the diet containing the Aroclor. Microscopically, centrolobular
    liver necrosis was found in chicks fed the first 2 compounds. Atrophy
    of the spleen and porphyria were observed in all dosed groups.

        Table 35.  Toxicity of PCBs for birds

    Species                 Sex      Age          Routea     PCB type          Parameter      Dose/concentration      Reference

    Bobwhite quail                   10 days      diet       Aroclor 1221      5-d LC50       >6000                   Hill et al. (1975)
    (Colinus virginianus)            10 days      diet       Aroclor 1232      5-d LC50       3002 (2577-3501)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1242      5-d LC50       2098 (1706-2610)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1248      5-d LC50       1175 (966-1440)         Hill et al. (1975)
                                     10 days      diet       Aroclor 1254      5-d LC50       604 (410-840)           Hill et al. (1975)
                                     10 days      diet       Aroclor 1260      5-d LC50       747 (577-937)           Hill et al. (1975)
                                     10 days      diet       Aroclor 1262      5-d LC50       871 (702-1069)          Hill et al. (1975)
                            male     1 year       oral       Aroclor 1268      acute LD50     >2000                   Hudson et al. (1984)

    Japanese quail                   14 days      diet       Aroclor 1221      5-d LC50       >5000                   Hill & Camardese (1986)
    (Coturnix coturnix               14 days      diet       Aroclor 1232      5-d LC50       >5000                   Hill & Camardese (1986)
    japonica)                        14 days      diet       Aroclor 1242      5-d LC50       >6000                   Hill & Camardese (1986)
                                     14 days      diet       Aroclor 1248      5-d LC50       4819 (4267-5443)        Hill & Camardese (1986)
                                     14 days      diet       Aroclor 1254      5-d LC50       2929 (2516-3409)        Hill & Camardese (1986)
                                     14 days      diet       Aroclor 1260      5-d LC50       2195 (1861-2589)        Hill & Camardese (1986)
                                     14 days      diet       Aroclor 1262      5-d LC50       2304 (1978-2684)        Hill & Camardese (1986)

    Table 35.  (cont'd).

    Species                 Sex      Age          Routea     PCB type          Parameter      Dose/concentration      Reference

    Mallard                          10 days      diet       Aroclor 1221      5-d LC50       >5000                   Hill et al. (1975)
    (Anas platyrhynchos)             10 days      diet       Aroclor 1232      5-d LC50       >6000                   Hill et al. (1975)
                                     10 days      diet       Aroclor 1242      5-d LC50       3182 (2613-3879)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1248      5-d LC50       2798 (2264-3422)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1254      5-d LC50       2699 (2159-3309)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1260      5-d LC50       1975 (1363-2749)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1262      5-d LC50       3008 (2461-3634)        Hill et al. (1975)
                            male     8-9 months   oral       Aroclor 1242      acute LD50     >2000                   Hudson et al. (1984)
                            male     8-9 months   oral       Aroclor 1254      acute LD50     >2000                   Hudson et al. (1984)
                            male     8-9 months   oral       Aroclor 1260      acute LD50     >2000                   Hudson et al. (1984)
                            male     8-9 months   oral       Aroclor 1268      acute LD50     >2000                   Hudson et al. (1984)

    Red-winged blackbird                          diet       Aroclor 1254      6-d LC50       1500                    Stickel et al. (1984)
    (Agelaius phoeniceus)

    Ring-necked pheasant             10 days      diet       Aroclor 1221      5-d LC50       >5000                   Hill et al. (1975)
    (Phasianus colchicus)            10 days      diet       Aroclor 1232      5-d LC50       3146 (2626-3948)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1242      5-d LC50       2078 (1843-3879)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1248      5-d LC50       1312 (1166-1477)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1254      5-d LC50       1091 (968-1228)         Hill et al. (1975)
                                     10 days      diet       Aroclor 1260      5-d LC50       1260 (1106-1433)        Hill et al. (1975)
                                     10 days      diet       Aroclor 1262      5-d LC50       1234 (1086-1402)        Hill et al. (1975)

    Table 35.  (cont'd).

    Species                 Sex      Age          Routea     PCB type          Parameter      Dose/concentration      Reference

    Starling                                      diet       Aroclor 1254      4-d LC50       1500                    Stickel et al. (1984)
    (Sturnus vulgaris)

    Brown-headed cowbird                          diet       Aroclor 1254      7-d LC50       1500                    Stickel et al. (1984)
    (Molothrus ater)

    Grackle                                       diet       Aroclor 1254      8-d LC50       1500                    Stickel et al. (1984)
    (Quiscalus quiscula)

    a  oral = acute oral test (result expressed as mg/kg body weight); diet = dietary test (result expressed as mg/kg diet).

    Miranda et al. (1987) studied the effects of acute oral exposure of
    Japanese quail to Aroclor 1242 (100, 250, or 500 mg/kg), or
    2,4,2',4'-tetrachlorobiphenyl and 3,4,3',4'-tetrachlorobiphenyl (both
    87.6 mg/kg) in corn oil. Control birds received only the corn oil. The
    birds were killed after 48 h. All the PCB compounds caused a
    significant increase in porphyrin content and delta- aminolevulinic
    acid synthetase (ALA-S) activity in the small intestine and liver. All
    the compounds increased the cytochrome P-450 content of the liver. In
    the intestine, the P-450 content was only increased by Aroclor 1242
    and 2,4,2',4'-tetrachlorobiphenyl. The activity of 7-ethoxyresorufin
     O-deethylase was increased by all compounds in both the intestines
    and liver. In the liver, 7-ethoxycoumarin  O-deethylase (ECOD)
    activity was unchanged or decreased, but, in the intestines, ECOD
    activity increased with dose. No tissue differences in ECOD activity
    were found after treatment with 2,4,2',4'-tetrachlorobiphenyl and
    3,4,3',4'-tetrachlorobiphenyl. It was concluded that the small
    intestine was more responsive than the liver to the porphyrinogenic
    effect of a single oral dose of PCBs, and, that the induction of drug
    metabolizing enzymes in the quail was tissue-specific, depending on
    the PCB preparation used.

    Day-old chicks were fed on a diet containing 500 mg PCBs/kg. All the
    birds died between the third and the tenth week of dosing; a reduction
    in the dosage to 250 mg/kg delayed the onset of death until the
    thirteenth week (Platonow & Funnell, 1971). Mortality did not occur in
    chickens dosed at 200 mg/kg over a period of 3 weeks (Flick et al.,
    1965). Harris & Rose (1972) fed one-day-old broiler chicks diets
    containing 100, 200, or 400 mg PCBs/kg (Aroclors 1242, 1254, and
    1260). No mortality occurred at doses of up to 100 mg/kg, over a
    period of 4 weeks. Over the same period, all the birds on Aroclor
    1242, 60% of the birds on Aroclor 1254, and none of the birds on
    Aroclor 1260, died, all at a dosage of 400 mg/kg. Holleman et al.
    (1976) fed day-old broiler chicks and turkey poults on diets
    containing Aroclor 1242 at 38, 75, or 150 mg/kg for 4 weeks. The
    authors found increased mortality at 75 mg/kg (21% mortality) with the
    chicks, but this was not significantly different from controls;
    however, there was a significant increase in deaths at 150 mg/kg (75%
    mortality). Significantly increased mortality was not found in the
    turkeys at any dose level, but the mortality rate in the controls was
    33%. Both the 75 and 150 mg/kg diets produced oedema and other lesions
    attributed to PCB toxicity. Prestt et al. (1970) maintained Bengalese
    finches on a diet containing various concentrations of Aroclor 1254.
    The estimated dose rate for 50% mortality, over 56 days, was 254 mg/kg
    per day.  Egg production

    Most studies demonstrating the lowering of egg production by PCBs were
    conducted on chickens. The most severe effects came from dosing with
    Aroclors in the middle of the range of chlorination (Aroclors
    1232-1254). The literature has been reviewed by Stendall (1976).

    Platonow & Reinhart (1973) fed chickens with Aroclor 1254 at either 5
    or 50 mg/kg diet over 39 weeks. Egg production was erratically reduced
    with the lower doses and sharply reduced on 50 mg/kg diet. A dietary
    dose of 2 mg/kg did not have any reproductive effects on chickens over
    9 weeks (Lillie et al., 1974) or after 39 weeks (Platonow & Reinhart,
    1973). Scott et al. (1975) showed a 10% reduction in egg production in
    chickens related to egg residues of PCBs (Aroclor 1248) of 3 mg/kg.
    When egg residues reached 4.5 mg/kg, the production rate was further
    reduced. A significant reduction in egg production was demonstrated by
    Call & Harrell (1974) after dosing Japanese quail with 3 different
    Aroclors at 62.5-5000 mg/kg, over 33-264 days.  Hatchability and embryotoxicity

    Aroclors reduced the hatchability of chicken eggs. In 2 studies,
    Lillie et al. (1974) examined the effects on hatchability of Aroclors
    1221, 1232, 1242, 1248, and 1268, all fed at 2 or 20 mg/kg diet, over
    9 weeks. Aroclors 1221 and 1268, with low and high chlorination,
    respectively, showed no effects at 20 mg/kg diet. Aroclor 1248
    produced some adult mortality at 20 mg/kg diet and nearly eliminated
    hatching of the eggs produced. Aroclor 1242 showed similar, but
    slightly less severe, effects; there was even less effect with Aroclor
    1232. Cecil et al. (1974) tested a similar range of PCBs at the same
    dosages (2 and 20 mg/kg diet). They also reported no effects for
    Aroclors 1221 and 1268. Aroclors 1254, 1232, 1242, and 1248 reduced
    hatchability, as in the previous study. PCBs that reduced hatchability
    also produced abnormalities in the chicks. The fertility of eggs was
    not affected by any of the treatments. Females were artificially
    inseminated with semen collected from males fed a similar diet of
    PCBs. Scott (1977) dosed chickens with 0.5, 1, 10, or 20 mg PCBs/kg
    diet and found no effect at the 2 lowest doses. Hatchability was
    reduced at 10-20 mg Aroclor 1248/kg diet. Kosutzky et al. (1979) dosed
    chickens with 2 other PCBs (Delor 103 and 105) (42 and 54%
    chlorination, respectively), at 5 mg/kg diet, for 6 weeks; there was
    little effect on hatchability, which returned to control levels soon
    after a return to a clean diet. Solomon et al. (1973) studied the
    effects of PCBs (Aroclor 1254) on pheasants. The birds were dosed at
    weekly intervals, for 17 weeks, with gelatin capsules containing
    50 mg/bird for the hens or 25 mg/bird for the cock birds. No effects
    were observed on fertility and there was no increase in the numbers of
    abnormal embryos. In another study, Ax & Hansen (1975) maintained
    white leghorn pullets on a diet containing Aroclor 1242 or 1254 at
    20 mg/kg, or 2,4,5,3',4'-pentachlorobiphenyl, for a period of 10
    weeks. Average embryonic mortality was found to be significantly
    increased, i.e., 54.7, 59.2, and 74% for the 3 compounds,
    respectively. In the same study, the authors found that average

    embryonic mortality in eggs laid by birds dosed with
    2,5,2'-trichloro-, 2,5,2',5'-tetrachloro-, or 2,4,5,2',4',5'-
    hexachlorobiphenyl was not significantly different from that in the
    controls. When both broiler breeder hens and leghorn hens were fed
    diets containing either 20 or 50 mg Aroclor 1242/kg, for 1 week, the
    hatchability of the eggs laid was reduced by 67.3 and 26.8%,
    respectively, of control levels, on the 50 mg/kg diet (Briggs &
    Harris, 1973). Hatchability also was reduced at 20 mg/kg diet, but the
    time required to achieve the same depression as that found with the
    higher dose was doubled. Even after dosing had finished,
    embryotoxicity continued and, in fact, increased until, after 6 weeks,
    hatchability was between 0 and 10% of controls for both birds at both

    Chickens given Aroclor 1254 at 50 mg/litre in the drinking-water, for
    6 weeks, showed progressive reduction in egg hatchability. This fell
    to zero after 3 weeks (Bush et al., 1974). Hatchability remained at
    almost zero for the first 8 weeks of the chickens receiving control
    water following dosing, but returned to normal after a further 8

    Platonow & Reinhart (1973) fed chickens on a diet containing 50 mg
    Aroclor 1254/kg for 39 weeks. There was some adult mortality; egg
    production and hatchability fell almost to zero. Residues of Aroclor
    in the last eggs produced ranged between 25 and 50 mg/kg. After 6
    weeks of uncontaminated food, egg residues dropped and hatchability
    improved. The authors reported that egg residues of less than 5 mg/kg
    had no effect on hatchability, whereas residues greater than
    10-15 mg/kg led to embryotoxic effects. Scott et al. (1975) related
    hatchability to egg residues. At residue levels of 3 mg/kg,
    hatchability was reduced by 44%, and, at residue levels of 4.5 mg/kg,
    it was reduced to almost zero.

    The yolk-sacs of eggs from pheasant, mallard, goldeneye duck, and
    black-headed gull were injected with 3,4,3',4'-tetrachlorobiphenyl at
    0.1 mg/kg (pheasant and mallard) or 1.0 mg/kg (pheasant, goldeneye
    duck and black-headed gull) after 4 or 5 days of incubation (Brunstrm
    & Reutergardh, 1986). A significant decrease in the hatching rate was
    seen only in pheasants, at the highest dose. At a dose of 1 mg/kg, all
    the embryos died before hatching, but, at 0.1 mg/kg, no effect on
    hatching was observed. No gross abnormalities were noted in either
    hatched chicks or dead embryos. A great difference was noted by the
    authors between the avian embryos in this study and chicken embryos,
    with regard to sensitivity towards tetrachlorobiphenyl. In chicken
    embryos, a dose of 0.004 mg/kg, administered on day 4 of incubation,
    gave a significant reduction in hatching. At 0.02 mg/kg, no embryos
    survived to hatching (Brunstrm & Darnerud, 1983). Carlson & Duby

    (1973) injected Aroclors directly into chicken eggs on the first day
    of incubation, or 9 days later. Aroclor 1242 severely limited
    hatchability at levels of more than 2.5 mg/kg. Aroclors 1254 and 1260
    had no effect at 10 mg/kg. Delaying the injection until day 9 of
    incubation reduced the effect of Aroclor 1242. With 5 mg/kg injected
    at day zero, hatchability was 8.3%; when the same dose was given at
    day 9, 82% of eggs hatched. These results are not compatible with
    those of Scott et al. (1971) who reported that most embryonic deaths
    occurred late in incubation. However, these authors were not dosing
    the eggs directly, but measuring the residues in eggs from dosed
    females. PCBs administered directly into the eggs may not produce
    effects comparable with those produced by the same material received
    from the mother hen, because of different distribution in the egg.
    Platonow & Reinhart (1973) dosed hens at 50 mg/kg diet. Early in the
    study, the majority of embryo deaths occurred late in incubation. As
    the study progressed, the time of embryo death moved to earlier in the
    incubation period. Bush et al. (1974) showed that there was greater
    mortality for any given egg residue as the period of dosing the mother
    hen progressed. Their experimental chickens were dosed for 6 weeks at
    50 mg/litre drinking-water and then kept for a further 20 weeks on
    clean water. On day 11 of the study, an egg yolk residue of 50 mg/kg
    was associated with 50% mortality in the embryos. On day 131, 50%
    mortality was associated with an egg residue of only 10 mg/kg. The
    greatest toxicity of PCBs for chicken embryos occurred after 11 weeks
    of clean water, that is 17 weeks into the study. At this stage, the
    eggs would be receiving doses of PCBs from material stored within the
    hen. Late in the study, residues of between 6 and 8 mg/kg in egg yolks
    (equivalent to 3.6 mg/kg whole egg) were correlated with between 14
    and 36% mortality. Platonow & Reinhart (1973) reported that egg
    residues greater than 10-15 mg/kg caused embryotoxic effects whereas
    low residues of less than 5 mg/kg did not produce any effects.
    Abnormalities were reported by Cecil et al. (1974) in 34% of 843
    embryos that died during their study. The most common abnormality was
    oedema, which was seen in 50% of all chicks showing any abnormality.
    Tumasonis et al. (1973) also reported deformities in chicks.  Eggshell thinning

    Since the 1950s, thin eggshells have been characteristic of many wild
    bird populations, though the effect was not noticed until some time
    afterwards. Thin shells has been a contributory factor to reduced
    reproductive capacity, particularly in birds of prey. The main
    chemical causing thin shells appears to be DDE, a metabolite of DDT.
    It has been shown to cause thin eggshells in laboratory experiments,
    as well as through the correlation of field data. Literature on thin
    eggshells has been reviewed by Cooke (1973). Most experimental studies
    using PCBs have shown no effect on shell thickness. Peakall (1971), in
    the first controlled study, dosed ring doves at 10 mg PCBs (Aroclor
    1254)/kg diet or at 25 mg (equivalent to 160 mg/kg body weight)

    injected ip. Shells were ashed and weighed. Two separate studies on
    dietary dosing showed no effect on shell weight. In the first study, 2
    groups of birds were compared, in the second the same birds were used,
    comparing their eggs before and after dosing. Injection of PCBs, 1-4
    days prior to egg laying, also had no effect. Studies on mallard dosed
    with Aroclor 1254 at 25 mg/kg diet and bobwhite quail dosed at
    50 mg/kg diet over 2 years and also on mallard dosed at up to
    500 mg/kg diet for 5 weeks (Heath et al., 1972) showed no effects on
    shell thickness. There was an apparent shell thickening of about 6% at
    the highest dose, which could not be statistically confirmed. The same
    authors outlined results from a study on white leghorn chickens.
    Aroclor 1242 at 10 or 100 mg/kg diet or Aroclor 1254 at 100 mg/kg diet
    did significantly reduce shell thickness. There were no measurable
    effects of: Aroclor 1242 at 1 mg/kg diet, Aroclor 1254 at 10 mg/kg, or
    Aroclor 1260 at 100 mg/kg diet (Heath et al., 1972). Experimental
    details and detailed results were not given for the work on chickens.
    Lillie et al. (1974) dosed chickens with a range of Aroclors, at 2 or
    25 mg/kg diet, for 9 weeks. Aroclors 1248 and 1242 greatly reduced the
    hatchability of eggs and caused some adult mortality, but failed to
    cause thinning of the eggshells. When Britton & Huston (1972) fed
    single comb White Leghorns Aroclor 1242 at 80 mg/kg diet, for 6 weeks,
    no effects on shell thickness were observed. Dahlgren & Linder (1971)
    failed to demonstrate any deleterious effects on the eggshells of
    pheasants, dosed by gelatin capsule, once a week for 17 weeks, with
    doses of Aroclor 1254 up to 50 mg. Call & Harrell (1974) fed various
    Aroclors in the diet to Japanese quail for 21 days. Very significant
    shell thinning was found with Aroclors 1254 and 1260 at doses of 1250
    and 1000 mg/kg diet, respectively. At these high doses, egg production
    was severely diminished and shell dimensions were based on very few
    eggs. Adult mortality might have occurred at these doses; the paper
    does not make it clear whether this actually happened. At lower doses
    of 78.1 and 62.5 mg/kg diet of Aroclors 1254 and 1260, respectively,
    there was also significant egg-shell thinning and reduced egg
    production. Aroclor 1242 was tested at 312.5 and 5000 mg/kg diet and
    both doses caused shell thinning, though this was to a lesser degree
    than with other Aroclors at similar doses. Risebrough & Anderson
    (1975) showed that eggshells thinned by dietary DDE were not further
    affected by adding PCB (Aroclor 1254) to the experimental DDE diet.
    Results on shell thinning are, therefore, not completely clear. It is
    generally agreed that PCBs do not affect birds in this way and the few
    results suggesting shell effects are regarded as anomalous or
    difficult to interpret, because of experimental design. Shell thinning
    can occur because of several different direct and indirect factors.
    DDE and sulfanilamide have direct effects on the deposition of calcium
    in the shell or on its mobilization from the skeleton, which acts as a
    calcium store. PCBs are more likely to affect shells indirectly by
    reducing food consumption; none of the studies cited above reported

    whether individual birds took less food because of the dosing with
    Aroclors. Haseltine & Prouty (1980) fed 24 pairs of mallard with
    Aroclor 1242 at 0 or 150 mg/kg diet for 12 weeks and reported a
    reduction in shell thickness of 8.9%. They pointed out that all
    females laying thin-shelled eggs showed a significant depression in
    body weight. This, they regarded as sufficient explanation for the
    shell thinning. Much of the shell thinning found in Japanese quail
    eggs, laid by females given a single oral dose of Aroclor 1254 of
    500 mg/kg body weight, was thought to be due to reduced food
    consumption (Haegele & Tucker, 1974).

    Biessmann (1982) did not find any effects on eggshell thickness on
    dosing Japanese quail with Clophen A60 at levels of up to 150 mg/kg
    diet. However, the breaking strength of the eggs was reduced.

    Hill et al. (1976) fed 6-month-old laying Japanese quail hens a diet
    containing 10 mg Aroclor 1242/kg, for 40 days. Eggs were collected and
    measured and, after 40 days, were found to have significantly thinner
    shells (5.2%) than the controls. The authors stated that the handling
    of the birds and the diet had no effect on food consumption or hen
    weights during the test. This is the only study showing shell thinning
    at moderate dose levels without an effect on food consumption. The
    question of whether PCBs can cause shell thinning, therefore, remains
    open.  Effects on the male

    Platonow & Funnell (1971) kept day-old, white leghorn cockerels on a
    diet containing 250 mg Aroclor 1254/kg, for up to 13 weeks. They found
    a significant reduction in the weight of both combs and testes after 9
    and 13 weeks and, a reduction in comb weight, only, after 6 weeks of
    dosing. In a later study, Platonow & Funnell (1972) found a more
    severe effect at 500 mg/kg; the comb was significantly reduced in
    weight after just one week of dosing and the testicular weight
    significantly reduced after 4 weeks, relative to the controls. The
    control combs and testes increased in weight during the course of the
    study; treated birds failed to develop either comb or testes.

    Lillie et al. (1974) did not find any effects on weight gain, food
    intake, or semen characteristics in leghorn cockerels fed Aroclor 1248
    at 10 or 20 mg/kg diet for 8 weeks. They also did not find any effects
    on fertility or hatchability of fertile eggs laid by similarly dosed
    females. Liver weights were significantly increased at both dose
    levels, and heart weights were significantly decreased at the highest
    dose.  The effects of stress

    Stress, imposed in various ways, increases the sensitivity of birds to
    PCBs. Stress seems to have its effect by increasing the mobilization
    of fat. Lower fat storage decreases the attenuation of PCB toxicity
    seen when fat uptake of the material acts as an effective temporary
    detoxification mechanism. Dahlgren et al. (1972) showed that brain
    residues were higher in pheasants subject to starvation stress than in
    unstressed birds dosed at the same rate. deFreitas et al. (1972)
    obtained similar results using cold stress or starvation in pigeons.
    As a corollary, biochemical adaptation to stress is inhibited by
    exposure to PCBs. This is presumed to be due to residues in non-lipid
    tissues (Dieter, 1974).  Physiological, biochemical, and behavioural effects

    Jefferies & Parslow (1972) dosed young lesser blackbacked gulls
     (Larus fuscus) with daily gelatin capsules containing Aroclor 1254
    at 50, 100, 200, or 400 mg/kg body weight, for 8 weeks. Mean
    individual thyroid weights were significantly increased by 32% (taking
    all dosed birds as a single group). There was also an increase in the
    mean cross-sectional area of the thyroid. There was, however, no
    dose-related effect of PCBs on thyroid weight. The same authors
    (Jefferies & Parslow, 1976) showed a similar effect of increased
    thyroid weight when they dosed guillemots  (Uria aalge) for 45 days
    with Aroclor 1254 at 12 or 25 mg/kg body weight. Hurst et al. (1974)
    also found a significant stimulation of thyroid growth after feeding
    bobwhite quail a diet containing 5, 50, or 500 mg Aroclor 1260/kg for
    4 months. Spear & Moon (1985) raised ring doves on either a low iodine
    or normal diet. Insufficient iodine caused thyroid hyperplasia. This
    hyperplasia was reversed within 7 days by a single dose of
    3,4,3',4'-tetrachlorobiphenyl at 60 mg/kg body weight. The PCB
    treatment also caused a significant decrease in core body temperature
    and serum total thyroxine (T4) and triiodothyronine (T3). No effect,
    other than decreased serum T3 and T4, was caused by dosing doves with
    PCBs on a diet containing normal iodine levels.

    Behavioural effects of PCBs have been noted by several authors.
    Peakall & Peakall (1973) reported decreased parental attentiveness in
    ring doves dosed at 10 mg Aroclor 1254/kg diet. Kreitzer & Heinz
    (1974) measured the avoidance response (from a moving silhouette) in
    Japanese quail chicks, for 14 days before, and 8 days after, dosing
    with Aroclor 1254 at 200 mg/kg diet. After dosing, the avoidance
    response was significantly reduced. Normal responsiveness to the
    silhouette was not recovered after 6 days on a clean diet. Two
    examples of hyperactivity in birds were also noted. European robins
    fed one mealworm/day containing 5 g Clophen A50, for 11-13 days,
    showed increased migratory restlessness (Ulfstrand et al., 1971).
    There were similar tendencies in redstart fed one mealworm/day,
    containing 11g Clophen A50, for 12 days, it was estimated that the
    birds had ingested 132 g of PCB overall (Karlsson et al., 1974).

    A reduction was reported by Dobson (1981)in the nest-building activity
    of pigeons  (Columba livia), dosed orally, by gelatin capsule, with
    15 mg Aroclor 1254/day, throughout a courtship cycle. The birds
    produced a nest but the number of twigs used was reduced compared with
    the controls. Reproductive and thyroid hormones were measured in blood
    plasma samples, taken each day during the courtship cycle. While the
    patterns of hormone secretion remained the same in both control and
    treated birds (rises and falls of hormone levels occurred at
    comparable times in the 2 groups) the absolute circulating levels of
    the hormones were changed by the treatment. Both thyroxine and
    luteinising hormone levels in the treated birds were elevated relative
    to the controls. The levels were significantly higher, except at the
    beginning and the end of the cycle. It was concluded that hormone
    levels were unaffected, except when they would naturally be changing,
    suggesting an interference with the feedback control of hormone
    secretion and a central nervous site of action. Tori & Peterle (1983)
    kept mourning doves  (Zenaida macroura carolinensis) on a diet
    containing Aroclor 1254 at 10 or 40 mg/kg for 42 days. The doves were
    then paired and observed each day for 30 days. Both treatments
    significantly increased the mean number of days in the courtship
    phase; only 4 out of 8 pairs on 10 mg PCBs/kg completed this phase and
    moved onto the nesting phase; none of the birds on 40 mg/kg had
    completed the courtship phase within 30 days. Behaviour was scored to
    measure intensity and, at both doses, this was significantly reduced
    overall. Although dosed birds formed pair-bonds approximately 4 days
    sooner than controls, there was no significant difference in the
    length of the pair-bond formation period or in behaviour scores during
    this period in doves fed 10 mg/kg. The length of time spent in the
    courtship period was extended significantly (by 8.5 days) by PCBs at
    10 mg/kg. Behaviour scores were not significantly affected, but dosed
    birds averaged 32% lower scores. Of the birds reaching the nesting
    phase, there was no significant difference between controls and dosed
    birds with regard to length of time spent nesting or behaviour scores
    in the nesting phase. PCBs, however, significantly delayed the onset
    of nest initiation (by approximately 7 days) and, therefore, egg

    Japanese quail were dosed with Clophen A60 in the feed at 150 mg/kg,
    from the first week of life up to 42 days of age, while the birds were
    developing sexually and becoming reproductively mature (Biessmann,
    1982). In females, progesterone levels were not greatly affected by
    the PCBs, but estradiol levels in blood plasma were lower before
    sexual maturity and were less stable during egg laying. In males,
    levels of testosterone and dihydrotestosterone (the primary
    metabolite) were not affected. Quail fed up to 150 mg Clophen A60/kg
    diet during the time of sexual maturation (second to fourth weeks of
    age) showed delayed onset of egglaying and a diminished capacity to
    lay eggs. Hormone levels in both males and females were not
    significantly different from those in the controls.

    A possible mechanism for central nervous effects was provided from
    studies on neurotransmitters. Dopamine and noradrenalin were depleted
    in the brain of the ring dove in a dose-related manner with increasing
    brain residues of PCBs (Heinz et al., 1980).  Interactive effects with other chemicals

    The only information on the interaction between PCBs and other
    chemicals in birds has shown PCBs to be additive and not synergistic.
    Kreitzer & Spann (1973) carried out tests on several pairs of
    chemicals to study pesticidal synergism in young pheasants and
    Japanese quail. Two PCBs were used in the study. Aroclor 1262 and
    malathion showed additive results, when fed in the diet to 16-day-old
    Japanese quail. Aroclor 1254 and DDE were also additive, when given in
    the diet to 9-day-old quail. Another study on the possible interactive
    effects of PCBs was conducted by Heath et al. (1972), who found that
    feeding Aroclor 1254 and DDE in the diet to 14-day-old Japanese quail
    gave additive results. There was no evidence of mutual potentiation or

    7.3.4  Terrestrial mammals

    Acute oral LD50 values reported for PCBs in mink ranged from >750 to
    4000 mg/kg body weight. Acute LD50 values for 3 Aroclors in mink were
    determined by Aulerich & Ringer (1977) after administration orally, by
    gavage, or by intraperitoneal injection. Mortality was assessed 4 days
    after i.p. administration and 14 days after oral administration. The
    lethality of the Aroclors was found to be inversely related to the
    chlorine content; Aroclor 1221 was most toxic and Aroclor 1254 least
    toxic (Table 36). This is in marked contrast to the situation in
    birds, where toxicity was correlated positively with chlorine content
    of the PCBs (see section 7.3.3).

    Table 36.  Acute toxicity of Aroclors for minka

    Aroclor             LD50 (mg/kg body weight)
                        Intraperitoneal              Oral

    Aroclor 1221        >500-<750                    >750-<1000
    Aroclor 1242         1000                        >3000
    Aroclor 1254        >1250-<2250                   4000

    a  From: Aulerich & Ringer (1977).  Short-term toxicity

    Ferrets  (Mustela putorius furo), fed a diet containing 20 mg of
    Aroclor 1242/kg, for 8 months, developed enlarged, thickened, and
    deformed toe-nails with hyperkeratosis at the junction of the skin and
    sponchium, and dysplasia of the root of the nail and the matrix. The
    same diet containing Aroclor 1016 did not produce these effects
    (Bleavins et al., 1982).

    Bleavins et al. (1980) showed the ferret to be less sensitive to PCBs
    than the mink, though LD50 values were not determined. Aroclor 1242
    at 20 mg/kg diet killed all mink (3 males and 12 females) to which it
    was fed. The same diet did not kill any ferrets, though it did cause
    reproductive failure.

    No mortality occurred in mink fed a diet containing 1 mg PCBs/kg over
    183 days (Wren et al., 1987a).

    Hornshaw et al. (1986) conducted 28-day LC50 tests on mink, using
    Aroclor 1254, in a study to investigate the effects of age, season,
    and diet on the toxicity of PCBs; no effects were noted on any of
    these parameters. In replicate tests, the calculated LC50 values
    varied between 79 and 84 mg/kg diet (48-132, range of confidence
    limits). The authors noted that the period of observation after dosing
    was critical in assessing the results, since mortality continued after
    dosing had stopped and the PCBs were persistent in the body. Taking
    total mortality over 28 days of dosing and a further 7 days of
    observation, the LC50 values fell to between 47 and 58 mg/kg diet.

    A consistent finding among various studies is an effect of PCBs on
    food consumption and, therefore, on body weight. The most detailed
    analysis of food consumption during the feeding of Aroclor 1254 to
    mink is presented by Hornshaw et al. (1986). Young mink fed the
    Aroclor over 28 days showed a dose-dependent decrease in the amount of
    food consumed. The cumulative weight of food consumed over 5 weeks
    (1 week predosing and 4 weeks dosing) for controls was 7574 g. This
    was reduced progressively with increasing dose of Aroclor 1254 at 10,
    18, 32.4, 58.3, and 105 mg/kg diet to 6447, 6153, 4816, 3556, and
    2723 g, respectively. This led to loss of original body weight over
    the study period rising to more than 40% at the highest dose. Similar
    effects were seen in adults of both sexes; females, with a smaller
    initial body weight, were more severely affected than males. This
    effect has implications for the interpretation of results of dietary
    toxicity tests. The effects seen are a combination of the direct toxic
    effects of the compound and the indirect effects of progressive
    starvation. The doses of PCBs to which the animals are exposed must
    also be calculated with reduced food intake in mind; apparent doses
    are higher than real exposure as dose rates increase.

    Organ weights (expressed as a percentage of brain weight) were
    unaffected by Aroclor 1254 at doses up to 105 mg/kg diet, with the
    exception of the heart and the adrenal glands. Heart weight was
    reduced in both adult and young mink fed Aroclor 1254 at 58 mg/kg diet
    or more; adrenal weights were increased by doses of 13 mg/kg diet or
    more (Hornshaw et al., 1986).

    Female minks received 0.1 or 0.5 mg of 3,4,5,3',4',5'-
    hexachlorobiphenyl/kg diet, or 2.5 or 5.0 mg of 2,4,6,2',4',6'- and
    2,3,6,2',3',6'-hexachlorobiphenyl/kg diet for 12.5-14.5 weeks. In both
    studies, 3,4,5,3',4',5'-hexachlorobiphenyl was the most toxic isomer,
    causing high mortality and reduced body weights (Aulerich et al.,

    Clark & Prouty (1977) fed female big brown bats  (Eptesicus fuscus)
    on a diet of meal-worms containing 10 mg Aroclor 1254/kg. After the
    feeding period of 54 days, the bats were starved to simulate loss of
    body fat during the period of migration (when the animals do not
    feed). Two out of 12 bats died. The brain residues of 20 mg PCBs/kg at
    the end of the study were considered to be sub-lethal, since no
    neurotoxic symptoms were observed before death. Reproductive effects

    Experimental investigations of mink reproduction in relation to
    environmental pollution were carried out as a result of the reduced
    reproductive success seen after feeding farm mink with fish from the
    Great Lakes. Early studies, therefore, involved the analysis of fish
    for pollutants and the experimental feeding of both the fish and of
    mixtures of chemicals contained in various fish in the Great Lakes.

    Aulerich & Ringer (1977) performed a comprehensive series of feeding
    studies using coho salmon from 2 of the Great Lakes (Michigan and
    Erie), other fish species from the same source, salmon from the west
    coast of the USA, and various combinations of organochlorine

    In their first study, ocean fish (perch or whiting) were used as
    control diets and the reproductive performances were compared of dosed
    and undosed female mink mated with undosed males. Lake Michigan coho
    salmon, as 30% of the diet, had the most severe effect on
    reproduction, i.e., total reproductive failure, as measured by numbers
    of live kits surviving 4 weeks after parturition. Coho salmon from
    Lake Erie also reduced reproductive success; 12 females produced only
    7 kits still alive 4 weeks after birth. Two other species of fish from
    Lake Michigan produced less severe effects, 5 kits being produced on a
    diet of 30% bloater chub and 15 kits, on a diet of yellow perch.
    Controls produced more than 40 kits over the same period. Kits

    produced on Lake Michigan or Lake Erie fish, other than salmon, and
    surviving to 4 weeks of age showed significantly lower body weights
    than the controls. The small numbers of surviving kits from mothers
    fed Lake Erie salmon also showed reduced body weight at 4 weeks of
    age. No kits were produced after feeding Lake Michigan salmon diets to
    females (Aulerich & Ringer, 1977).

    In a long-term, low-level feeding study, 4 different Aroclors (1016,
    1221, 1242, and 1254) were included in the diet of mink at a rate of
    2 mg/kg. Groups of 8 female and 2 male animals were given this, or a
    control, diet for 11 months, from August to June. The reproductive
    performance of the animals on different diets is summarized, together
    with mortality, in Table 37. Body weight, haemoglobin levels, and
    haematocrit were monitored at monthly intervals during the study and
    no significant effects of PCBs were noted. Only one of the test diets
    (containing Aroclor 1254) adversely affected reproduction, with only 1
    live birth in the study period. This single kit was considerably
    lighter at birth than the controls and failed to survive 4 weeks after
    birth (Aulerich & Ringer, 1977).

    The reproductive effects of either Lake Michigan coho salmon or
    Aroclor 1254 were reversible, when animals were transferred to a
    control diet. Eleven females fed salmon as 30% of the diet for a year,
    and then given control food for a further year, produced young with an
    average litter size of 3.5 kits per mated female, in the second year
    of the study. No young had been produced in the first year of the
    study, during dosing. Similarly, 3 females given a year of control
    food following a year on a diet containing 5 mg Aroclor 1254/kg,
    produced an average of 4.3 young per mated female in the second year
    of observation (Aulerich & Ringer, 1977).

    In a later study, Bleavins et al. (1980) fed 2 different Aroclors at
    various dose levels to mink and ferrets. Aroclor 1242 was fed to mink
    at doses ranging from 5 to 40 mg/kg diet; Aroclor 1016 was given at
    only 20 mg/kg diet. Results for mortality and reproductive effects are
    summarized in Table 38, together with some data for Aroclor 1254 taken
    from Aulerich & Ringer (1977).

    There was a clear reproductive effect of Aroclor 1242 at a dose of
    5 mg/kg diet, but no significant mortality. Aroclor 1242 caused 66%
    mortality at 10 mg/kg diet and 100% mortality at 20 mg/kg diet.
    Aroclor 1016, at 20 mg/kg diet, caused some deaths of adults and
    reduced birth-weight and survival of kits, but the reproductive
    effects were considerably less severe than those caused by Aroclor
    1242 at 5 mg/kg diet. The authors (Bleavins et al., 1980) calculated
    dietary LC50 values for Aroclors 1242 and 1254, using their own data
    and data from Aulerich & Ringer (1977), to be 8.6 and 6.7 mg/kg diet,
    respectively. It is clear that reproductive effects are less marked at
    lower levels of chlorination of Aroclors.

        Table 37.  Effects of Aroclors on mortality and reproduction in minka

    Aroclorb     Adult females            Kits
                 Number        Number     Number     Number born              Whelped/      Alive at      Average weight
                 died (%)      mated      whelped                             female        4             (g  SE) at
                                                     live         dead        mated         weeks         birth

    Control        0           8          8           28          5           4.1            18           9.9  0.32

    1016           0           8          8           28          8           4.5            16           9.2  0.33

    1221          12           7          7           43          1           6.3            37           9.6  0.22

    1242          12           7          7           35          4           5.6            32           9.3  0.27

    1254          12           7          2            1          1           0.3             0           5.4

    a  From: Aulerich & Ringer (1977).
    b  Aroclors all fed at 2 mg/kg diet.

    Reproductive effects on mink were reported by Jensen et al. (1977),
    who fed groups of 10 females at 0.05, 3.3, or 11 mg PCBs/kg diet (type
    unspecified), for 66 days. The highest dose eliminated successful
    reproduction, whereas 3.3 mg/kg severely reduced the number of kits
    born per female.

        Table 38.  Summary of mortality and reproduction in mink fed various dietary
               levels of Aroclors

    Treatment level       Period fed       Number dead/      Number of kits/
    (mg/kg diet)          (days)           total number      female

    Aroclor 1254a
        0                 280               1/7              5.0
        0                 297               0/8              4.1
        2                 297               1/8              0.3
        5                 280               2/7              0.0
       10                 280               5/7              0.0

    Aroclor 1242b
        0                 247               3/30             4.9
        2                 297               1/8              5.6
        5                 247               1/15             0.0
       10                 247              10/15             0.0
       20                 192c             15/15             0.0
       40                 138d             15/15             0.0

    Aroclor 1016b
        0                 247               3/30             4.9
        2                 297               0/8              4.5
       20                 247               3/15             6.3

    a  From Aulerich & Ringer (1977).
    b  From: Bleavins et al. (1980).
    c  All mink died within 192 days on diet.
    d  All mink died within 138 days on diet: no females survived to whelping.

    Wren et al. (1987b) did not find any significant effects on numbers of
    kits produced or surviving to weaning age (5 weeks) after feeding mink
    with Aroclor 1254 at 1.0 mg/kg diet, over 183 days. The dosing period
    covered the seasonal period when the animals came into breeding
    condition as well as a period of giving birth to the young. Although
    the weights of kits born to dosed females were not significantly
    different at 1 week postpartum, the weight gain of kits was then
    affected and weights were significantly different from those of the
    controls at ages 3 and 5 weeks. At age 5 weeks, when the kits were
    weaned, the mean body weight of kits of the controls was 227.8 g,
    while the mean body weight of kits fed by dosed mothers was 161.2 g.

    Similar effects on the reproduction of female mink fed PCBs (type
    unspecified) were observed by Jensen et al. (1977), i.e., a reduction
    in the numbers of whelps born per pregnant female. The authors killed
    the females after they had given birth and examined the numbers of
    implantation sites in the uterus. This did not differ statistically
    between groups (on average, 6.6 in control females, 6.1, in females
    fed 5 mg/kg diet, and 4.5, in females fed 15 mg/kg diet PCBs).
    However, the number of kits born to the same females showed a marked
    effect of the PCBs: 5.1 (on average) born to control mothers; 2.9, to
    mothers fed 5 mg/kg diet PCBs, and 0, to mothers fed 15 mg PCBs/kg
    diet. The authors concluded that the effects of PCBs occur at the time
    of implantation or later, causing resorption of implanted embryos.

    Male mink seemed unaffected by doses of Aroclor that caused
    reproductive effects in females. Males, dosed and mated with undosed
    females, fathered normal numbers of kits (Aulerich & Ringer, 1977).
    Wren et al. (1987b) did not note any effects of Aroclor 1254, fed to
    male mink over 183 days at a rate of 1 mg/kg diet. Testicular size and
    testicular histology were unaffected by the PCBs at any stage of the
    reproductive cycle.

    Treatment of adult, male, white-footed and white mice with Aroclor
    1254 in the diet at a level of 400 or 200 mg/kg (equivalent to 57 or
    29 mg/kg body weight), for 2 weeks resulted in a reduced testicular
    spermatozoan concentration and, in the white-looted mice, a reduced
    absolute weight of the seminal vesicles. In both strains, the absolute
    weights of the testes and the final body weights were unchanged
    (Sanders & Kirkpatrick, 1975; Sanders et al., 1977).

    The reproductive performance of 27 pairs of white-looted mice (44-222
    days of age), was compared with that of 26 control pairs, within 60
    days of exposure to Aroclor 1254 at a dietary level of 200 mg/kg.
    One-third of the exposed pairs did not survive the exposure but
    produced at least one litter. The number of pairs producing at least
    one litter was reduced by 65% and the number producing 2 or more
    litters was reduced by 91%. Litter size was not affected. No offspring
    survived to weaning in the 7 first litters of PCB-fed pairs (Merson &
    Kirkpatrick, 1976).

    A group of 10 pairs of wild-caught, white-footed, mice and groups of
    18 (12 weeks of age) and 19 (16 weeks of age) pairs of
    laboratory-raised, white-looted mice received Aroclor 1254 in the diet
    at 10 mg/kg. Control groups comprised 10, 15, and 20 pairs,
    respectively. The reproductive performance of the first group was
    recorded for 18 months. The duration of the other studies varied from
    7 to 15 months. The number of young per litter, 28 days after birth,
    was lower in all treated groups. In laboratory-raised mice paired at
    12 weeks of age, the birth interval was increased and the number of
    young per litter at birth reduced (Linzey, 1987a). The second
    generation of mice, maintained on the same diet as their parents, did
    not differ in weight at birth, but were significantly smaller at 4, 8,
    and 12 weeks of age. A similar trend was observed in the few young of
    the third generation. The uterus, ovaries, and accessory glands, but
    not the testes, weighed less in exposed groups than in the controls
    (Linzey, 1987b). Linzey (1987a) reported similar reproductive effects
    of Aroclor 1254 at a much lower dose.

    The author suggested that the major consistent effect on the survival
    of the young being fed milk was the result of much higher levels of
    PCBs being transported via lactation than via the placenta.

    Cottontail rabbits  (Sylvilagus floridanus) were fed Aroclor 1254 at
    10 mg/kg diet for 12 weeks, and then transferred to a clean diet and
    allowed to breed. No effects were observed on any reproductive
    parameters, and reduction in food availability did not change this
    lack of effect (Zepp & Kirkpatrick, 1976).  Physiological effects

    Wren et al. (1987a) examined histologically various organs in male and
    female mink, dosed for 183 days with Aroclor 1254 in the diet. At
    autopsy, no effects were seen on the histology of the pituitary and
    adrenal glands. Brain histology also appeared normal. Thyroid
    follicles gave the general appearance of minimal activity but did not
    differ between treatments. Measurement of plasma thyroid hormones
    (T3 -triiodothyronine; T4 -thyroxine) did not show any significant
    differences between treatments in male mink. Females showed reduced
    circulating T3 in a single sample, in January, but no other
    differences were seen at other stages of the study. Thyroxine levels
    were not affected at any time.

    7.4  Effects on organisms in the field

    The acute toxicity of PCBs is relatively low for most species and will
    not, therefore, kill enough individuals to affect populations.
    However, because of the high potential for bioaccumulation sufficient
    residues of PCBs may build up to cause direct lethal effects over
    time. Although PCBs are almost universally present in the tissues of

    organisms in the environment, there are relatively few examples of
    proved effects of these residues on populations of the organisms.
    Sublethal effects, affecting populations by reducing reproduction or
    growth, are possible, but difficult to prove, because PCBs are always
    present with other environmental contaminants. Many possible effects
    of PCBs in the environment have been suggested in the literature, but
    few have actually been investigated in the field. It has proved
    difficult, if not impossible, to relate residues of PCBs in tissues to
    possible sublethal toxic effects; residues found after laboratory
    dosing cannot be directly related to the field situation.

    7.4.1  Plants

    Klekowski (1982) studied the ostrich fern  (Matreuccia struthiopteris)
    growing in the flood plain of the Housatonic River, Massachusetts,
    USA. This area of the river is contaminated with PCBs from land-fill
    sites containing waste materials from the manufacture of transformers
    in the nearby city of Pittsfield. Contamination with PCBs (principally
    Aroclor 1254) had been a regular feature of the river area for a
    period of more than 40 years. The frequency of somatic mutations in
    the fern population was compared to a control population from an
    uncontaminated area. The levels of PCBs in river sediments ranged from
    1.4 to 139 mg/kg dry weight; at the site where the majority of fern
    spores were collected, the level of PCBs was 26.3 mg/kg. The somatic
    mutation frequency for the contaminated population was 5.2-6.2 times
    higher than that for the controls. It is not known whether similar
    genetic damage had occurred in other inhabitants of this contaminated
    habitat. No other studies seem to have been conducted on the possible
    effects of PCBs, from land-fill sites, on plants.

    7.4.2  Fish

    There have been many suggestions in the literature that PCBs might
    affect populations of fish in the wild. Studies attempting to
    demonstrate such an effect are few and, generally, inconclusive or

    Olofsson & Lindahl (1979) used the ability of the cod  (Gadus morrhua)
    to react to different velocities of water under rotary flow, to
    examine the effects of water pollution. Cod sampled from polluted
    waters off the Swedish Coast were compared with cod sampled from
    unpolluted areas. The ability of the fish to react to rotary flow was
    significantly reduced in animals from polluted areas. However, the
    authors were unable to relate the reduced reaction of the fish to

    levels of various pollutants measured in muscle tissue. Experimental
    studies showed that PCBs affected the reactions of the fish; the
    residue level in muscle that was associated with this effect was
    1.8 mg/kg. This was 30 times greater than the actual residues of PCBs
    measured in the cod from the polluted area. While the authors stated
    that the distribution of the PCBs in experimental fish and fish taken
    from the wild would almost certainly have been different, that PCBs
    exert the above effect in the field must be regarded as not proved.

    Zitko & Saunders (1979) collected eggs of Atlantic salmon  (Salmo
     salar) from various areas and measured the PCB contents of the eggs
    that proved infertile. The hatchability of different batches of eggs
    was tested in the laboratory. No correlation was found between
    residues of PCBs and the hatchability of the eggs; in fact the batch
    of eggs showing the lowest hatchability also showed very low residues
    of PCBs. However, it should be stated that hatchability was seldom
    affected by PCBs in laboratory experiments; effects were more usually
    seen on the developing young. Hogan & Brauhn (1975) related the
    survival of fry hatched from rainbow trout  (Salmo gairdneri) eggs to
    the contamination of the eggs with PCBs. Five batches of eggs hatched
    in 1971 had shown percentage mortalities, 30 days after hatching,
    ranging from 10 to 28%. The eggs contained total organochlorine
    residues of between 0.31 and 1.30 mg/kg, the majority of which was
    PCBs. A batch of eggs collected in 1972 showed 75% mortality, 30 days
    after hatching. Many of the fish hatched with such deformities as
    scoliosis, lordosis, kyphosis, absence of caudal vertebrae, cranial
    deformities, and projecting mandibles. This batch of eggs contained
    2.7 mg PCBs/kg and 0.09 mg total DDT/kg (metabolites present not

    Westin et al. (1983) investigated the effects of PCBs, passed on to
    the eggs of striped bass  (Morone saxatilis) by the female fish, and
    of PCBs in food organisms fed to the hatched larvae. The hatchlings
    were fed on brine shrimp ( Artemia sp.) from 2 sources, one
    contaminated with PCBs and the other not. No effects of either
    maternal PCBs or PCBs from the food were found. Residues of PCBs in
    young fish decreased consistently throughout the study, which was
    conducted in water free of PCBs. The authors suggested that PCBs from
    the mother and from food are unlikely to affect the offspring in the
    wild. A similar conclusion was drawn about the survival of lake trout
     (Salvelinus namaycush) fry, hatched from eggs contaminated with PCBs
    in the wild (Willford, 1980). Eggs taken from the wild and hatched in
    clean water showed good survival of the offspring (suggesting that
    residues of PCBs in the eggs were not responsible for the failure of
    the species in Lake Michigan). However, experimental exposure of eggs,
    and hatched larvae/fry to levels of PCBs in water and food, similar to

    those found in the lake, led to high mortality and the increased
    occurrence of deformities in fry. The author concluded that the levels
    of PCBs in lake water and food items would be sufficient to lead to
    the population decline seen in the lake. It should, however, be
    pointed out that other factors in Lake Michigan could also have
    contributed to the failure of the lake trout. Sea lampreys had
    increased in number in the lake and could have caused population
    decline by predation.

    The relationships between the occurrence of hepatic diseases and
    specific chemicals present in sediment were studied by Malins et al.
    (1987). The concentrations of PCBs in the sediments from 4 urban and 2
    non-urban areas in Puget Sound USA, were determined. In 3 of the
    sites, the concentrations were <0.01 and, in the other 3, the
    concentrations ranged from 0.11 to 0.53 mg/kg. Over 900 individual
    organic compounds were found. To study the fish diseases, English sole
     (Parophrys vetulus), rock sole  (Lepidopsetta bilineata), and
    Pacific staghorn sculpin  (Leptocottus armatus) were collected. The
    organs of fish containing the greatest number of lesions were the
    liver, kidneys, and gills. Liver cell adenoma, hepatocellular
    carcinoma, cholangiocellular carcinoma, haemangioma, and fibroma
    constituted major types of liver lesions. Statistically significant
    correlations between levels of chemicals in sediment and hepatic
    neoplasms in the bottom-dwelling fish suggest a general
    cause-and-effect relationship, but there is little firm evidence about
    the actual cause of these neoplasms, in particular, in this case,
    whether PCBs were involved.

    7.4.3  Birds

    In studies on chickens and different PCB-mixtures, Vos et al. (1970)
    and Vos & Koeman (1970) found that the induction of subcutaneous and
    abdominal oedema, centrilobular liver necrosis, hydropericardium, and
    higher mortality was more or less related to the presence of
    tetrachloro- and pentachlorodibenzofurans, as impurities in the PCB

    From the middle of February to the end of March 1968, an epizootic
    disease closely resembling chicken oedema disease occurred in Japan.
    Two million chickens were involved, of which 400 000 (20%) died. The
    clinical signs were laboured breathing, droopiness, ruffled feathers,
    high mortality, and decreased egg production. Autopsy revealed marked
    subcutaneous oedema, hydropericardium, ascites, pulmonary oedema,
    muscular ecchymosis in the thorax or inside of the thigh, and
    yellowish mottled appearance of the liver. The cause of the disease
    was found to be Kanechlor 400 contamination of the feed. Experimental
    reproduction of these symptoms with Kanechlor 400 was successful. The
    remaining sample chicken feed contained 1300 mg of Kanechlor 400/kg
    feed (Kuratsune et al., 1972).

    Administration of PCBs leads to an atrophy of lymphoid tissue in
    chickens (Flick et al., 1965; Vos & Koeman, 1970), and in pheasants
    (Dahlgren et al., 1972). Vos & de Roij (1972) and Vos & van
    Driel-Grootenhuis (1972) came to the conclusion that these effects
    could be attributed to an immunosuppressive effect of PCBs. Vos & de
    Roij (1972) suggested that the ability of PCBs to increase the
    susceptibility of ducklings to duck hepatitis virus (Friend & Trainer,
    1970) and of fish to fungal disease (Hansen et al., 1971) could be
    attributed to this immunosuppressive effect.

    Koeman et al. (1973) analysed 6 adult cormorants  (Phalacrocorax carbo
     sinenis), found dead in the wild, for PCB residues. Three additional
    birds were shot and 6 fully grown nestlings were also taken. The
    authors also carried out a study on 5 cormorants that were dosed with
    the PCB, Clophen A60 (a 60% chlorinated PCB mixture that seemed to
    correspond most closely with the PCBs profile of material found in
    dead birds taken from the wild). The birds were initially dosed with
    the PCBs in the diet, but, subsequently, the PCBs were administered
    orally in gelatin capsules, until they died. PCB levels in the brains
    and livers of the dead birds found in the wild were higher, overall,
    than the levels obtained by dosing. The authors thought it highly
    probable that this was indicative of PCB poisoning of the birds in the
    wild. Without further detailed studies, this conclusion can only be
    implied. The higher residues in dead birds from the wild suggest that
    a large body burden was taken up, relatively safely in fat, and
    released quickly on starvation, prior to death.

    A field study, to show the effects of a sub-lethal dose of PCBs on
    puffin  (Fratercula arctica) breeding success and survival, was
    conducted by Harris & Osborn (1981). A total of 150 puffins, trapped
    on the Isle-of-May National Nature Reserve, Fife, Scotland, were
    implanted, 108 with between 30 and 35 mg of PCBs (Aroclor 1254) and 42
    with sucrose as controls. The test chemicals were implanted in
    open-ended silastic tubes into the peritoneum. All birds were then
    marked and released back into the wild. The same birds returned in
    successive years to breed on the island. Breeding and survival of the
    implanted birds were monitored through the breeding seasons of 1977,
    1978, and 1979. Observations on survival were also made in 1980. Some
    implanted birds were killed on recapture, and analysed for PCBs, in
    each of the years of observation. No effects on survival or breeding
    were seen. PCB levels in fat increased by a factor of between 10 and
    14 compared with levels in birds that had not been implanted with the

    Herring gull  (Larus argentatus) reproductive success in the area of
    the Great Lakes declined with increasing residues of organochlorines
    in the birds. Breeding success improved as these residues fell in the
    1970s (Weseloh et al., 1979). The species was chosen as an indicator
    of environmental pollution in the area and has been extensively
    studied. Poor nesting success in the species is related to high
    embryonic mortality (Gilbertson & Hale, 1974). Abnormal chicks have
    been reported for herring gulls and also for other fish-eating species
    from the area including: night herons, ring-billed gulls, common
    terns, and Caspian terns (Gilbertson et al., 1976). Eggs from these
    species contained residues of PCBs, but it was not possible to
    directly relate these residues to chick abnormalities. Gilbertson &
    Fox (1977) found a correlation between total organochlorine content
    and the hatchability of herring gull eggs. While it cannot be shown
    that PCBs were directly responsible for the overall effect, PCB
    residues in the livers of hatched chicks, were the only contaminant
    that was significantly correlated with the presence of pericardial
    oedema. Weseloh et al. (1979) concluded that the reproductive success
    of the herring gull can only safely be correlated with total
    organochlorine residues and that the possible effects of PCBs cannot
    be isolated. Shell thinning usually attributed to DDE alone, does not
    occur significantly in these gulls. Hays & Risebrough (1972) suggested
    that PCB residues in terns were responsible for the high percentage of
    abnormal young in a colony from Long Island Sound.

    Twenty-four black-crowned night heron  (Nycticorax nycticorax) eggs
    were collected at the San Francisco Bay National Wildlife Refuge in
    1983. Twelve of these were collected from separate nests, when
    late-stage embryos were pipping and an additional egg was randomly
    collected from each nest for organochlorine analysis. Other anomalies
    and skeletal defects were not apparent. Embryonic weights (with
    partially absorbed yolk sacs removed) were 15% lower in comparison
    with controls collected at Patuxent Wildlife Research Center.
    Crown-rump length and femur length were shorter in the San Francisco
    Bay embryos. The geometric mean PCB concentration was 4.1 mg/kg wet
    weight with a range of 0.8-52.0 mg/kg. A negative correlation existed
    between embryonic weight and log-transformed PCB residues in whole
    eggs, suggesting a possible impact of PCBs on embryonic growth. DDE
    did not show such a correlation (Hoffman et al., 1986).

    Klaas & Swineford (1976) found low PCB residues (0.26-3.4 mg/kg wet
    weight) in 35 screech owl  (Otus asio) eggs (16 of which were known
    to be addled) taken from the wild; there was no relationship between
    the presence of residues and hatching failure. Dosing captive screech
    owls with 3 mg Aroclor 1248/kg (McLane & Hughes, 1980) showed no
    effects on eggshell thickness, number of eggs produced, young hatched,
    or young fledged. Residues of PCBs measured in the eggs ranged between
    3.9 and 17.8 mg/kg. This range covers the egg residue levels that were
    clearly associated with effects in experimental chickens and

    Eggs from 315 clutches of sparrowhawks  (Accipiter nisus) from 9
    sites in Scotland were examined by Newton & Bogan (1978) and Newton
    (1979). Eggs that failed to hatch were collected and analysed for
    PCBs, DDE, and for aldrin and dieldrin. Statistical analysis showed
    little variation in residues within a single clutch, but wide
    variation between clutches, even clutches from the same female in the
    same area in different years. On this basis, analysis of single eggs
    was taken as representative of the complete clutch. Analysis of
    variance and multiple regression were used to try to relate particular
    effects, such as shell thinning, addling of eggs, and late-embryo
    mortality, with particular organochlorines. Results showed that
    addling of eggs was significantly correlated with levels of both DDE
    and PCBs. Because of the close correlation between residues of DDE and
    PCBs in eggs of sparrowhawks, it is not possible to tell whether only
    one, or both, organochlorines were involved in egg addling. PCBs
    showed the strongest relationship with addling. Shell thinning was
    more directly linked with DDE. Newton et al. (1982) concluded that the
    level of PCB residues in merlin  (Falco columbarius) was not
    sufficiently high to have exerted any effect on the breeding success
    of this species in Britain. Hodson (1975) did not find any effects of
    PCB residues in Richardson's merlin in Canada and attributed all
    effects to DDE. A relationship between DDE residues and shell thinning
    and the breeding success of the bald eagle  (Haliaetus leucocephalus)
    was demonstrated by Wiemeyer et al. (1984). PCB residues were
    correlated with DDE residues. The authors concluded that contaminants
    other than DDE contributed no more than a minor role in reproductive
    effects on this species in the field. It was considered that PCBs
    might have contributed to reproductive problems, but that the evidence
    was unclear. Newton et al. (1986) presented a statistical analysis of
    residues in sparrowhawks related to reproductive success. Study
    populations from many sites throughout the British Isles showed
    different reproductive success. Some of this variation could be
    related to levels of organochlorine compounds in eggs. When shell
    thickness index was related to PCBs alone, a significant negative
    correlation was detected. However, once DDE residues were included,
    PCBs did not improve the model. It was concluded that DDE alone could
    account for all of the reproductive effects recorded and that PCBs, at
    the levels of contamination found, probably did not play a role in
    reduced breeding performance. Newton et al. (1988) examined new data
    on residues of organochlorine compounds in the eggs of the peregrine
    falcon  (Falco peregrinus) and also reassessed older data. Total
    information considered included information on residues and breeding
    success covering the period 1963-86. The PCB residues found were not
    considered to have had a significant effect on breeding success in
    this species. In earlier studies, Wiemeyer et al. (1975, 1987) could
    not isolate any single organochlorine compound as responsible for the
    reduced breeding success of the fish-eating osprey  (Pandion

    Heinz et al. (1983) examined the breeding success of groups of
    red-breasted mergansers  (Mergus serrator) on islands in northwestern
    Lake Michigan. This is a fish-eating species and, since eating fish
    from the Great Lakes had been shown to affect captive mink, the
    authors investigated the possible effects of contaminants (measured in
    blood samples taken from the birds) on reproduction. Other,
    non-contaminant effects were also examined. Many organochlorine
    contaminants of the environment were found in eggs including 14
    different organochlorine compounds; there was a correlation between
    the levels of many of these compounds. No relationship could be
    established between levels of PCBs, or any other contaminant, and the
    breeding success of the birds. The hatching success of dabbling ducks
    also seemed to be largely unaffected by feeding in Lake Michigan,
    despite the presence of residues of PCBs and other contaminants in the
    eggs (Haseltine et al., 1981).


    Effects of PCBs have been shown in laboratory studies on many domestic
    species of birds, but few wild species. However, there are many
    studies showing measurable residues of PCBs in wild birds. PCBs have
    been implicated in population decline in several bird species in
    different parts of the world, but it is seldom easy to demonstrate
    directly the effects of PCBs on populations of birds in the field.
    This is largely because PCB residues invariably occur together with
    other organochlorine compounds, such as DDE and dieldrin. Separating
    out individual effects can only be done completely satisfactorily when
    laboratory data are available for the same species and each individual
    pollutant. Some idea of likely effects can be obtained from
    statistical manipulation of data from the field, though, in this case,
    correlation is the best that can be achieved. In practice, few field
    studies give the results that have been predicted from laboratory
    studies on other species; that is direct extrapolation from laboratory
    to field and species to species is not straightforward.

    7.4.4  Mammals

    Laboratory studies involving the dosing of non-laboratory mammals with
    PCBs are limited to a few species. Field studies have been conducted
    on a wider range of species.

    Norstrm et al. (1988) studied liver and adipose tissue specimens of
    polar bears (121 samples) obtained by Inuit hunters from 12 zones in
    the Canadian Arctic Archipelago in 1982-84 (see section 5.1.4). Six
    PCB congeners (Nos 99, 153, 138, 180, 170, and 194) constituted
    approximately 93% of total PCBs. The major PCBs accumulated belong to
    the group formed by combinations of 2,4-dichloro-, 2,4,5-,
    2,3,4-trichloro-, and 2,3,4,5-tetrachloro- substitution on each ring.

    Congeners with 2,4-, 3,4-dichloro-, 2,3,4-, 2,3,5-trichloro-,
    2,3,4,6-, 2,3,5,6-tetrachloro- or 2,3,4,5,6-pentachloro- substitution
    on one ring, such as PCB Nos 118, 138, 187, 183 and 196, which usually
    bioaccumulate readily in mammals, were all at lower levels compared
    with congeners with only 2,4,5-trichloro- and 2,3,4,5-tetrachloro-
    substitution in ringed seals, such as PCB Nos 153, 180, and 170. The
    last 3 congeners accounted for 71% of total PCBs in polar bears. Thus,
    it appears that the polar bear is able to metabolize PCB congeners in
    which there are nonchlorinated  para positions, adjacent
    nonchlorinated  ortho-meta positions, or both  ortho positions
    chlorinated in one ring.

    DeLong et al. (1973) measured various organochlorine pollutants (DDT,
    DDD, DDE, dieldrin, and PCBs) in sea lions from the Channel Islands
    off the Californian coast. Previous observations had recorded a high
    incidence of premature births in the population and organochlorine
    residues from females showing premature or normally-timed parturition
    were compared. Both DDT and PCB residues (measured against a standard
    of Aroclor 1254) were significantly higher in the females producing
    premature pups. Residues were estimated in the blubber, liver, and
    brain, and, for the first 2 tissues, the ranges of values, for the 2
    groups of females, did not overlap. Previous studies on both
    substances indicated possible reproductive effects in mammals,
    corresponding to some extent with those observed in the sea lions.
    Cause and effect could not, therefore, be directly established for
    each of the pollutants alone. Helle et al. (1976a) collected ringed
    seals from Simo, on the northern Bothnian Bay area of the Baltic Sea,
    in October/November. This population of seals shows reduced
    reproductive capacity and the sampled population showed only 27% of
    females pregnant, compared with other reports of 62.5% and 85-90% for
    the same species elsewhere. Residues of DDT and PCBs were compared
    with those in seals from other areas of the Baltic and found to be
    lower (Table 39). The seals sampled from Simo were divided into
    pregnant and non-pregnant groups (n = 15 and 26, respectively).

    DDT and PCBs levels were both significantly higher in non-pregnant
    animals (Table 40). All females from both groups showed a corpus
    luteum in one ovary, indicating that all had ovulated. About half of
    the non-pregnant females showed indications of an embryo having been
    implanted, which had subsequently aborted or resorbed. Again cause and
    effect could not be established directly. The authors pointed out that
    normally breeding Californian sea lions had DDT levels as high as
    those in Baltic seals showing reproductive failure and proposed this
    as an indication that PCBs were the causative agent.

        Table 39.  Levels of DDT and PCBs (mean mg/kg  S.E.) in extractable fat of
               blubber from ringed seal from the Baltic Seaa

    Area                         Number          DDT            PCBs

    Northern most part           40              110  10b      69  4.4b
    of the Bothnian Bay

    Gulf of Bothnia              33              200  28       110  15

    a  Data from: Helle et al. (1976a).
    b  Value significantly different from those in the Gulf of Bothnia (P < 0.05).

    Table 40.  Levels of DDT and PCBs (mean mg/kg  S.E.) in extractable
               fat of blubber in non-pregnant and pregnant ringed seal of
               reproductive agea

                       Number          DDT                PCBs

    Non-pregnant       26              130  13b          77  5.2b
    Pregnant           15              75  11            56  6

    a  Data from: Helle et al. (1976a).
    b  Values for non-pregnant females significantly higher than
       those of pregnant females. DDT P < 0.01; PCBs P < 0.05.

    It was further pointed out that the group of non-pregnant females
    would include some animals that were not pregnant for reasons other
    than the presence of organochlorine compounds. In a later paper (Helle
    et al., 1976b), the same authors reported that some of the
    non-pregnant females showed abnormal uteri with stenosis or occlusion
    of the uterine horns. They, therefore, subdivided the non-pregnant
    females in their second sample into those showing the anatomical
    abnormality and those not. Both DDT and PCBs were significantly higher
    in the occluded group than in pregnant animals. Non-pregnant females,
    without occlusions, showed residues of DDT and PCBs not significantly
    different from those in pregnant animals, and significantly lower than
    those females with occlusions. Residues in males were comparable with
    the highest residues found in females (Table 41).

    There is some indication of a positive correlation between PCB
    residues and age in male seals but not in females (Addison et al.,
    1973; Addison & Smith, 1974). These observations are presumed to
    indicate that females excrete some of their body burden of PCBs in the

    These field observations have been confirmed in a study in which
    captive seals were fed on fish from the Wadden Sea (where reproductive
    problems had been found in the wild seal population) and compared with
    controls fed fish from the Atlantic Ocean. Seals eating the
    contaminated fish, which differed from the control fish only in the
    PCB content, showed the same failure to carry the fetus successfully
    to term seen in the wild (Reijnders, 1986; see section 7.2.5).

        Table 41.  Levels of DDT and PCBs (mean mg/kg  S.E.) in extractable fat from
               blubber of ringed seal from Simo, Bothnian Baya

    Group     Sample description       Number       DDT             PCBs

    I         pregnant females           24          88  9.7        73  6.6
    II        non-pregnant females       29         130  10        110  7.8
              with stenoses and
    III       non-pregnant females       8          100  15         89  11
              with normal uteri
    IV        fetuses                    24          62  4.3        49  3.0
    V         males                      24         130  18        100  13

    probability of similarity between test groups (t-test)

    I-II                                            P < 0.01b       P < 0.01b
    I-III                                           P > 0.05        P > 0.05
    I-IV                                            P < 0.05c       P < 0.01b

    a  Data from: Helle et al. (1976b).
    b  Groups significantly different at the 99% level.
    c  Groups significantly different at the 95% level.

    Subramanian et al. (1987) collected samples of blubber from male
    Dall's porpoise  (Phocoenoides dalli) in the northwestern Pacific
    Ocean and analysed the samples for organochlorine content. Blood
    samples taken from the same animals were analysed for testosterone, a
    male sex steroid, and for aldosterone, another steroid hormone,
    responsible for blood electrolyte balance. Testosterone levels in
    blood were correlated with blubber levels of DDT and PCBs; as the
    organochlorine content increased, testosterone levels decreased. There
    was no relationship between blubber organochlorine levels and blood
    levels of aldosterone. The number of samples was small (n = 12) and,
    though the relationship between DDT and testosterone was significant,
    the apparent relationship with PCBs was not. The animals were sampled
    outside the breeding season, when testosterone levels would be
    expected to be low.

    In a study by Clark & Lamont (1976), 26 pregnant big brown bats
     (Eptesicus fuscus) were collected from the field and maintained on a
    control diet, in captivity, until they gave birth to young. Levels of
    PCBs were measured in both the adults and the offspring. The
    concentrations of PCBs in litters with dead young were significantly
    greater than in litters where both young were born alive (mean for
    litters with dead young was 2.44 mg/kg; mean for all other litters was
    0.34 mg/kg). The contents of PCBs ranged between 1.07 and 1.96 mg/kg
    wet weight in adults and between 0.28 and 1.69 mg/kg in the young.  Appraisal

    PCBs have been implicated in population declines of seals and sea
    lions. Population decreases and reproductive failure have been
    observed in seals from the Baltic Sea, the Wadden Sea (southeastern
    North Sea), and the Gulf of St. Lawrence and in sea lions off the
    Californian coast. Poor reproduction has been correlated with residues
    of PCBs in the affected animals. A major problem in such studies is
    the occurrence of more than one chemical pollutant in the animals.
    PCBs are found together with other organochlorine compounds and heavy
    metals. Conclusions, therefore, have often relied on making the best
    correlation between observed effects and residues and checking cause
    and effect relationships on animals more amenable to laboratory study.
    A study on captive seals confirmed field observations on the effects
    of PCBs on these marine mammals (see section 7.2.5).


    Evaluation of the toxicity of Aroclors and other commercial PCB
    mixtures is complicated by numerous factors, including isomer and
    congener composition, differences in species susceptibility,
    quantitatively inconsistent data, and various degrees of contamination
    with toxic compounds, such as chlorinated dibenzofurans. Because of
    these factors and a lack of data for some of the Aroclors (most
    studies have been conducted on the higher chlorinated Aroclors), it is
    assumed that effects resulting from exposure to a specific Aroclor are
    representative of effects that may be produced by the other Aroclors.
    Many of the literature sources do not give details of the composition
    of the PCB mixture used in the studies.

    Reviews on the literature concerning the toxicity of PCBs are given by
    ATSDR (1989), Kimbrough (1980, 1987), Safe (1984), NIH (1985), and
    Lorenz & Neumeier (1983).

    8.1  Single exposures

    8.1.1  Oral

    (a) Aroclors

    The acute oral LD50 values for a number of Aroclors are presented in
    Table 42. The lowest oral LD50 in the rat was 1.0 g/kg body weight
    for Aroclor 1254 (Garthoff et al., 1981).

    In Sherman rats, lethal doses of Aroclor 1254 and 1260 in peanut oil
    caused ulceration in the duodenum and glandular stomach (Kimbrough et
    al., 1972). Osborne-Mendel rats, receiving a single oral dose of
    50 mg/kg body weight of Aroclor 1254 in corn oil, showed an increased
    relative liver weight together with the induction of microsomal
    enzymes, within 12 h (Litterst et al., 1974).

    Monkeys, sacrificed 4 days after gastric intubation of Aroclor 1248 at
    1.5 or 3.0 g/kg body weight (no vehicle reported), exhibited enlarged
    livers with proliferation of the endoplasmic reticulum and hypertrophy
    and hyperplasia of the gastric mucosa (Allen et al., 1974a).

    (b) Individual congeners

    Estimated oral LD50 values in 30 days for individual congeners in
    corn oil in Hartley guinea-pigs (probably single application) were
    0.5 mg/kg body weight for 3,4,5,3',4',5'-hexachlorobiphenyl, less than
    1 mg/kg body weight for 3,4,3',4',-tetrachlorobiphenyl, more than
    3 mg/kg body weight for 2,3,4,5,3',4',5'-heptachlorobiphenyl, and more
    than 10 mg/kg body weight for 2,4,5,2',4',5'-hexachlorobiphenyl
    (McKinney et al., 1985).

        Table 42.  Acute oral LD50s of Aroclorsa

    Aroclor    Species/strain        Sex/ageb           LD50 (g/kg body weight)             References

    1254       rat/Wistar            male/30 days       1.3                                 Grant & Phillips (1974)
                                     female/30 days     1.4
                                     male/60 days       1.4
                                     female/60 days     1.4
                                     male/120 days      2.0
                                     female/120 days    2.5
               rat/Sherman           male/weanling      1.295                               Linder et al. (1974);
                                     NR/adult           4-10                                Kimbrough et al. (1972);
               rat/Osborne-Mendel    male/adult         1.01 (single dose)                  Garthoff et al. (1981)
                                                        1.53 (5 doses over 2 1/2 weeks)
                                                        1.99 (5 doses, 1 day/week)

    1221       rat/Sherman           female/NR          4.0                                 Nelson et al. (1972)

    1260       rat/Sherman           NR/adult           4-10                                Linder et al. (1974)
                                     M/weanling         1.315

    1242       rat/Sprague-Dawley    male/adult         4.25                                Bruckner et al. (1973)

    1262       rat                   NR                 11.3                                Panel on Hazardous Trace
                                                                                            Substances (1972)

    a  References: ATSDR (1989); WHO (1976).
    b  NR = not reported.
       Information on solvents used was not available.

    8.1.2   Inhalation

    No acute data were available.

    8.1.3   Dermal

    Median lethal doses for a single application of Aroclors to the skin
    of rabbits ranged from >0.79 to <1.27 g/kg body weight for Aroclors
    1242 and 1248 in 50% corn oil, to >1.0 to <3.17 g/kg body weight for
    undiluted Aroclor 1221, and >1.26 to <2.0 g/kg body weight for
    Aroclor 1260 (Fishbein, 1974).

    8.1.4   Other routes

    (a) Aroclors

    With a single intravenous dose of Aroclor 1254 in a 1% lecithin-saline
    suspension, the LD50 for adult female Sherman rats was 0.4 g/kg body
    weight (Linder et al., 1974). The LD50s for Aroclor 1254 in DMSO for
    various mouse strains, following a single intraperitoneal injection,
    varied between 0.9 and 1.2 g/kg body weight (Lewin et al., 1972).

    (b) Individual congeners

    The intraperitoneal LD50 for 2,4,3',4'-tetrachlorobiphenyl in CF-1
    mice was 2.15 g/kg body weight, while that of its main metabolite,
    (5-hydroxy-2,4,3',4'-tetrachlorobiphenyl) was 0.43 g/kg body weight,
    which suggests that the acute toxicity of 2,4,3',4'-tetra-
    chlorobiphenyl might be attributable to this phenolic metabolite
    (Yamamoto & Yoshimura, 1973).

    In male Wistar and Charles-River CD rats, a single intraperitoneal
    dose of 1 mg 3,4,5,3',4',-pentachlorobiphenyl/kg body weight or a
    single oral dose of 50 mg 3,4,5,3',4',5'-hexachlorobiphenyl/kg body
    weight caused significant liver enlargement accompanied by fatty
    changes, atrophy of the thymus, and decreased relative spleen weights.
    Mono- ortho substituted biphenyls, such as 2,3,4,5,3',4'-hexa-
    chlorobiphenyl and 2,4,5,3',4'-pentachlorobiphenyl, induced these
    effects in the liver and thymus to a minor degree following an
    intraperitoneal dose of 50 mg/kg body weight, while di- ortho
    substituted hexachloro- and heptachlorobiphenyls did not cause adverse
    effects at this dose level (Kohli et al., 1979; Yoshihara et al.,

    8.2  Short-term exposures

    8.2.1  Oral  Aroclors

    (a) Mouse

    While oral intoxications by PCB mixtures in monkeys are easily
    recognizable by their effects on the skin, skin lesions in female ddN
    mice were only observed after 2-3 months of daily oral exposure to
    1.6 mg of a technical PCB mixture (48% chlorine) per mouse (80 mg/kg
    body weight per day) in olive oil. The lesions included alopecia,
    erosions, ulcerations, and eczematous changes around the eyelids. No
    increase in mortality and only slight growth retardation were seen
    after 26 weeks of exposure (Nishizumi, 1970).

    Koller (1977) found histological changes in the liver of mice exposed
    to 37.5 mg Aroclor 1254/kg diet for 6 months. A dose level of
    3.75 mg/kg diet was without effect.

    The oral toxicities of Kanechlor 400, 500, and 600 were compared by
    feeding male mice diets containing 300 mg PCBs/kg for 14 weeks
    (Kawanishi et al., 1975). Fatty degeneration and accumulation of
    pigment in the liver were observed in the mice fed Kanechlor 600.

    (b) Rat

    At lethal, oral doses of undiluted Aroclor 1242 (100 mg/kg body
    weight, every 2 days, for 3 weeks), Sprague-Dawley rats showed reduced
    body weight, thymus atrophy, chromodacryorrhea, progressive
    dehydration, and central nervous system depression with terminal
    ataxia and coma. Fatty changes were observed in the liver and kidneys
    (Bruckner et al., 1973).

    Male and female Sherman rats (10 animals of each sex) were fed diets
    containing 0, 20, 100, 500, or 1000 mg of Aroclor 1260 or Aroclor
    1254/kg (equivalent to 0, 1.5, 7, 36, or 72 mg/kg body weight,
    respectively) for 8 months. The animals receiving the 2 highest dose
    levels showed reduced growth. Female rats fed Aroclor 1260 at 500 and
    1000 mg/kg diet showed a high mortality rate. However, only 3 rats fed
    Aroclor 1254 at 500 mg/kg diet died. Significant increases in relative
    liver/body weight ratios were found for both Aroclors at all doses
    tested. Microscopically enlarged hepatocytes, cytoplasmic inclusions,
    increased lipid levels, and foamy cytoplasm were all found
    consistently. Adenofibrosis was found at higher doses (500 and
    1000 mg/kg diet) and corresponded to the glistering white areas seen
    on gross inspection. These areas also showed cholangiofibrosis
    (according to Kimbrough, synonymous for bile duct proliferation, bile
    duct adenomatosis, and fibroadenoma) (Kimbrough et al., 1972).

    Growth and mortality rates were unaffected in male Sprague-Dawley rats
    exposed for 1 year to a diet containing Aroclors 1248, 1254, or 1262
    at 100 mg/kg diet (equivalent to 5 mg/kg body weight) compared with
    controls (Allen et al., 1976).

    Female Charles-River CD rats were fed for 20 weeks on a diet
    containing 0, 10, 30, or 100 mg Aroclor 1254/kg (equivalent to 0, 0.5,
    1.5, and 5 mg/kg body weight, respectively). No increase in mortality
    rates occurred but growth inhibition was seen at 30 mg/kg from month 2
    onwards and at 100 mg/kg diet from week 2 onwards. Skin lesions,
    initially on the ears, were found after 10-20 weeks of exposure to
    Aroclor 1254 at all dose levels. The lesions, including alopecia and
    reddened and thickened skin with hyperkeratosis, subcutaneous oedema,
    and infiltration by polymorphonuclear leukocytes, also involved the
    nose, tail, and feet at the highest exposure level (Zinkl, 1977).

    Effects on the liver following exposure to PCB mixtures have mainly
    been investigated in rats (see section

    (c) Rabbit

    Four groups of 5 male and 5 female New Zealand White rabbits were
    administered 300 mg Aroclor 1221, 1242, or 1254 in corn oil, via a
    stomach tube, once a week, for 14 weeks. The fourth group received
    only the vehicle.

    The mortality rate was increased and the body weight gain was reduced
    after 14 weeks of oral exposure to Aroclor 1254, but not after
    exposure to Aroclor 1221 or 1242. Liver/body weight ratio and SGOT and
    SGPT activities were increased in the animals treated with Aroclor
    1242 or 1254, but not in those treated with Aroclor 1221. In the
    animals treated with Aroclor 1254, the smooth endoplasmic reticulum
    was condensed in the liver cell and formed hyalin inclusions which
    might have been accompanied by a loss of enzyme activity. Lipid
    accumulation, pigment deposition, nuclear changes, and necrosis were
    also found (Koller & Zinkl, 1973).

    (d) Pig

    Pigs, fed Aroclor 1242 or 1254 at a dietary level of 20 mg/kg
    (equivalent to 0.8 mg/kg body weight) for 91 days, showed gastric
    lesions, consisting of erosions and necrosis. Two pigs fed a high-dose
    regimen of 100 mg of Aroclor 1254/kg body weight for 11 days, also
    showed hypertrophy and hyperplasia of the gastric mucosa; this was
    also found in monkeys at low exposure levels (Hansen et al., 1976b).

    (e) Cow

    Holstein cows were studied throughout a complete lactation period, a
    non-lactating period, and 42 days of a subsequent lactation period for
    overt responses to Aroclor 1254. Four cows received daily doses of 0,
    10, 100, or 1000 mg/cow (1000 mg/cow is equivalent to 1.67 mg/kg body
    weight) over 60 days. The mean daily milk production and net energy of
    a complete lactation period did not differ between control and
    PCB-treated animals. At the end of the study, the concentrations of
    PCBs were 0.005, 0.021, 0.14 in blood plasma; 1.9, 10.9, 91.3 in milk
    fat, and 1.4, 6.9, 70.0 g/kg in adipose tissue for the 10, 100, and
    1000 mg PCB groups, respectively. No signs of impaired health,
    productivity, or changes in blood and urine chemistry were observed
    (Willett et al., 1987).

    (f) Monkey

    Rhesus monkey

    (i) Aroclor 1242

    Becker et al. (1979) exposed 6, 7-8-month-old, male Rhesus monkeys to
    Aroclor 1242 at levels of 0, 3, 10, 30, or 100 mg/kg diet (equivalent
    to 0, 0.12, 0.4, 0.4, 1.2, or 4.0 mg/kg body weight, respectively) to
    study changes in the stomach mucosa. At 3, 10, 30, and 100 mg/kg diet,
    4 monkeys died after 245, 146, 92, and 137 days of exposure,
    respectively. All exposed monkeys showed a decreased body weight gain.
    At all dose levels, changes were observed in the folds of the
    suborbital facial skin, and eyelids became swollen and red. From day
    71 at 3 mg/kg diet, days 69 and 77 at 10 mg/kg diet, and day 12 at the
    higher exposure levels, stomach biopsies revealed an apparent arrest
    of the differentiation of generative cells of the isthmus and neck
    regions into parietal and zymogenic cells. Mature parietal and
    zymogenic cells, which were found only in the bases of the glands,
    showed signs of injury, such as dilatation of the rough endoplasmic
    reticulum on the zymogenic cells, irregularity of the mitochondria in
    parietal cells, and irregular luminal membranes and an increase in the
    number of autophagic vesicles in both type of cells. The severity of
    the lesions was directly correlated with both duration and level
    exposure. A no-effect level was not obtained in this study.

    A group of 5 Rhesus monkeys, 1-2.5 years old, was exposed to a diet
    containing 1 mg of Aroclor 1242/kg (equivalent to 0.04 mg/kg body
    weight) for 133 days. A control group contained 4 monkeys. No adverse
    effects were found (McNulty et al., 1980).

    Characteristic lesions, metaplasia in epithelial structures, such as
    sebaceous glands, nail beds, gastric mucosa, and ameloblast
    surrounding unerupted teeth were reported to have developed in Rhesus
    monkeys, 13 months after a 40-day diet containing 400 mg Aroclor
    1242/kg (equivalent to 16 mg/kg body weight) (McNulty, 1985).

    (ii) Aroclor 1248

    A group of 5 (4 males and 1 female), 1-month-old Rhesus monkeys,
    without previous exposure, were administered 30 daily doses of 35 mg
    of Aroclor 1248/kg body weight, by gavage, in corn oil. Four animals
    received only corn oil. No mortality or clinical signs were observed,
    except for a decrease in body weight gain, slightly reduced food
    consumption, and anaemia. Mild microscopic changes were observed in
    the thymus, bone marrow, eye, skin, stomach mucosa, and liver
    (Abrahamson & Allen, 1973).

    Six male Rhesus monkeys, aged 1 1/2-2 years were fed a diet containing
    300 mg Aroclor 1248/kg for 3 months. Three animals served as controls.
    A decrease in body weight was observed. Within one month, all the
    animals fed PCBs had alopecia, subcutaneous oedema, particularly of
    the face, which manifested as swollen eyelids, erythema, and acneiform
    lesions involving the areas devoid of hair (Allen & Norback, 1972 cf.
    Hayes (1987)).

    Six adult, female Rhesus monkeys were fed 25 mg Aroclor 1248/kg diet
    for 2 months. Facial oedema, alopecia, and acneiform changes developed
    within 1 month and 1 animal died 2 months after removal from the
    experimental diet. In addition to the above changes, this animal
    showed anaemia, hypoproteinaemia, hypertrophy, and hyperplasia with
    invasion through the muscularis mucosa, focal haemorrhages and
    ulcerations of the gastric mucosa, and bone marrow hypoplasia. Eight
    months later, the 5 surviving animals continued to show clinical signs
    of intoxication. The PCB concentration in body fat, which, after 2
    months of treatment, averaged 127 mg/kg, had decreased 8 months later
    to 34 mg/kg (Allen et al., 1974a).

    Groups of 9 adult, female Rhesus monkeys (weight approximately
    5.6 kg), were fed a diet containing 0, 2.5, or 5.0 mg Aroclor 1248/kg
    (equivalent to 0, 0.1, or 0.2 mg/kg body weight, respectively) for an
    average period of 18.2 months. An additional group of 4 males was fed
    5.0 mg/kg diet. Control groups contained 12 female and 6 male monkeys.
    The Aroclor 1248 contained 4.4-8.7 ng polychlorinated dibenzofurans/kg
    diet. The female monkeys were mated with control males after 6 months
    of exposure (the reproductive effects are discussed in section Levels of Aroclor 1248 in adipose tissue reached a plateau
    after 1 year at 2.5 mg/kg diet and after 6 months at 5 mg/kg diet.
    Exposed males showed slight to moderate periorbital oedema and
    congestion of the eyes. Females showed an average 15% loss in body
    weight over the first 5 months of both exposures, while food
    consumption was normal. At 6 months, they all showed loss of hair,
    acne of the face and neck, and erythema and swelling of the eyelids.

    Skin biopsies revealed keratinization of the affected hair follicles.
    One female monkey died after 173 days of exposure to 2.5 mg/kg diet
    and one female died after 310 days of exposure to 5.0 mg/kg diet. Both
    animals developed terminal enteritis due to Shigellosis, which was
    resistant to treatment. At autopsy, the animals showed generalized
    alopecia, subcutaneous oedema, and acne. Microscopically, follicular
    epithelial hyperplasia with inflammation of the surrounding tissue,
    and keratinization of the hair follicles were observed. The livers
    showed focal areas of necrosis, enlarged hepatocytes, and fatty
    changes (Barsotti et al., 1976).

    The surviving females in the former study, 8 per dose level, were
    placed on a control diet for approximately 1 year after an average
    total intake of 270 and 498 mg Aroclor 1248, respectively. There was a
    gradual improvement in their physical condition and, within one year,
    their gross appearance was no longer different from that of the
    controls. However, their breeding performance was still affected, as
    outlined in section (Allen et al., 1980).

    (iii) Aroclor 1254

    Several pilot studies were carried out before this major study to
    characterize the toxicity of Aroclor 1254 and to compare the toxic
    findings in the Rhesus and Cynomolgus monkey (Tryphonas et al., 1984,
    1986b; see section In these pilot studies, Mes & Marchand
    (1987) and Mes et al. (1989a) measured the concentrations of PCBs in
    the blood, adipose tissue, and faeces.

    A preliminary report describes the results of an ongoing study after
    54 weeks of daily oral administration of gelatin capsules containing
    0, 5, 20, 40, or 80 g Aroclor 1254/kg body weight, in corn
    oil-glycerol, to groups of 16 adult, female Rhesus monkeys  (Macaca
     mulatta). At this stage, a slight decrease in body weight gain was
    observed in the exposed monkeys together with a decrease in water
    consumption, but not in food consumption. In week 52, the incidences
    of prominent nail beds, of nails separated from the beds, and of
    prominent tarsal glands increased in the animals exposed at 80 g/kg
    body weight (Arnold et al., 1984).

    Aroclor 1254, at a dose level equivalent to 280 g/kg body weight was
    given for 5 days per week to Rhesus monkeys  (Macaca mulatta) over a
    period of 27-28 months. Four animals were treated as described and 4
    animals served as controls. The Aroclor 1254 was administered in
    apple-juice-gelatin-corn oil emulsion. The weight of the monkeys at
    the beginning of the study was approximately 4 kg. Terminal clinical
    signs of varying severity included finger nail detachment, exuberant
    nail beds, weight loss, stomatitis, and normocytic anaemia. At
    necropsy, the bone marrow was hypocellular with cytoplasmic vacuoles
    in erythroid precursor cells. Histopathological lesions included

    dilatation of the tarsal gland ducts, atrophy, or absence of, splenic
    and lymphonodal germinal centres, bone marrow depletion, gingival
    erosion and ulceration, moderate mucinous hypertrophic gastropathy
    with cystic dilatation of occasional gastric glands, hepatocellular
    enlargement and necrosis, hypertrophy of biliary duct epithelium,
    hyperplasia of biliary ducts, hypertrophy of the gall bladder
    epithelium, and an equivocal increase in the number of lysosomes in
    the thyroid follicular epithelial cells. The terminal PCBs
    concentrations in a number of organs were as follows: adipose tissue
    106.7-2073.2 mg/kg (in control animals 0.26-0.65 mg/kg); brain,
    31.2-252.0 mg/kg; kidneys, 85.4-964.1 mg/kg; and liver,
    255.1-828.1 mg/kg tissue. The PCB concentrations were expressed on a
    mg/kg fat basis. It was concluded that skin appendicular lesions are
    good clinical indices of PCB exposure in monkeys and that
    lymphoreticular lesions (atrophy and absence of lymphoid follicular
    centres) are good indicators of impending or active immunological
    crisis (Tryphonas et al., 1986a).

    (iv) Miscellaneous studies

    Accidental exposure of a colony of 256 Rhesus monkeys to PCBs in a
    concrete sealant produced a disease characterized by high mortality,
    gradual weight loss, behavioural changes, alopecia, acne, facial
    oedema, swollen eyelids, diarrhoea, anaemia, poor breeding
    performance, and high incidences of abortions and still births.
    Samples of the concrete slabs within several buildings were obtained
    and analysed. Significant levels of PCBs (5280 mg PCB/kg sample) were
    found (Altman et al., 1979; McConnell et al., 1979).

    The effects of exposure to PCBs on the eyes were investigated by
    Ohnishi & Kohno (1979), who administered a banana injected with 0.5 mg
    of PCBs/kg body weight, daily, to 8 adult Rhesus monkeys of both sexes
    for 1-5 months. Two out of this group were fed PCBs with
    polychlorinated dibenzofurans (2.5 g/kg body weight). Four untreated
    animals were used as controls. One month after the onset of the
    exposures there was a reduction in body weight. When pressure was
    applied to the eyelids of treated monkeys, white secretions extruded.
    Within 3 months alopecia, swelling of the eyelids, and acne-form
    eruptions developed. The retina and choroid were normal. The
    histopathological changes in the eyelids were comparable in both
    groups of exposed monkeys and included the appearance of keratinous
    cysts and atrophy of the Meibomian glands with hyperkeratosis and
    hyperplasia of the ductal epithelium.

    Cynomolgus monkey

    Groups of 5 or 6 adult, female Cynomolgus monkeys  (Macaca
     fascicularis) were exposed to 3, weekly doses of 4.7 mg of Aroclor
    1248/kg body weight (equivalent to 2 mg/kg per day) or to 3-weekly
    doses of 11.7 mg of Aroclor 1254/kg body weight (equivalent to 5 mg/kg
    per day) in an apple juice-corn oil emulsion. The monkeys were exposed
    until necropsy at day 30-164 in dead or moribund condition. A control
    group contained 5 monkeys. In both exposed groups, body weight loss,
    emaciation, facial oedema, finger-nail loss, and lacrimation were
    observed. Common histopathological lesions were: dilated Meibomian
    gland ducts, mucinous hyperplasia and hypertrophy of the gastric
    mucosa, enlargement, fatty degeneration, and necrosis of hepatocytes,
    bile duct and gall bladder epithelial cell hypertrophy and
    hyperplasia, and thyroid changes in follicular cell size and the
    number of intracytoplasmic lysozomes.

    The onset of the signs and lesions of toxicity was not as rapid and
    uniform as that in Rhesus monkeys. Aroclor 1248 appeared more toxic
    than Aroclor 1254 (Tryphonas et al., 1984).

    Groups of 4 adult, Rhesus and 4 adult, Cynomolgus female monkeys,
    weighing 3.2-4.5 and 3.2-5.2 kg, respectively, received doses of
    Aroclor 1254 at 0 or 280 g/kg body weight for 5 days/week (equivalent
    to 200 g/kg per day) in an apple-juice-gelatin-corn oil emulsion for
    27-28 and 12-13 months, respectively. This study showed that the
    Rhesus monkey is more susceptible to PCB toxicity than the Cynomolgus
    monkey (Tryphonas et al., 1986b).

    Hori et al. (1982) exposed 1 female, Cynomolgus monkey to daily doses
    of Kanechlor 400 (without detectable quantities of polychlorinated
    dibenzofurans) at 2 mg/kg body weight, in olive oil. They also exposed
    3 monkeys to a PCB mixture with a chromatographic pattern similar to
    that of the Yusho mixture (2 or 4 mg/kg body weight), and 1 monkey to
    2 mg/kg body weight of the same mixture, without detectable quantities
    of polychlorinated dibenzofurans (detection limit not given). The
    doses were administered 6 times per week in a piece of banana. Two
    controls received only the vehicle. The 2 monkeys receiving the Yusho
    mixture at 4 mg/kg body weight died within 4 and 8 weeks,
    respectively. The other monkeys were kept for 20 weeks. In all monkeys
    exposed to the Yusho mixture, toxic effects were similar to those
    already described above. In addition, there was cytoplasmic
    vacuolation and dilatation of the convoluted tubules with cytoplasmic
    casts in the kidneys. The other 2 mixtures induced less severe
    reductions in body weight gain, immunosuppression and
    histopathological alterations in liver, kidneys, and periorbital skin.
    The effects in the animals fed a diet with dibenzofurans yielded
    enhanced decreases in body weight, immunosuppression, fatty liver and
    histological changes, in addition to hair loss, acne-form eruptions,
    oedema of the eyelids and cornification of the skin, compared with the
    other test substances.  Individual congeners

    (a) Monkey

    Rhesus monkeys have been exposed to various congeners, as outlined in
    Table 43.

        Table 43.  Toxicity of PCB-congeners in Rhesus monkeysa

    Congenerb              No. of       Exposure               Time clinical    Deaths
                           animals/                            toxicity
                           group        mg/kg      period      first noted
                                        diet       (days)      (day)

    2,5,4'-TriCB           4            5          84          -                0

    3,4,3',4'-TCB          3            3=>1c      215         14-21            3
                           5            1          38          27               1

    2,5,2',5'-TCB          3            3=>1c      215         -                0
                           5            1=>5d      200         -                0

    3,4,5,3',4',5'-HCB     1            0.1        127         117              1
                           4            0.5        63          28-30            4
                           1            1          57          20               1

    2,4,5,2',4',5'-HCB     4            15         122         -                0
                           1            65         63          -                0

    2,4,6,2',4',6'-HCB     4            15         122         -                0
                           1            65         64          -                0

    2,3,6,2',3',6'-HCB     4            15         122         -                0

    a  From: McNulty et al. (1980); McNulty (1985); Iatropoulos et al. (1977).
    b  TriCB = trichlorobiphenyl; TCB = tetrachlorobiphenyl; HCB = hexachlorobiphenyl.
    c  Dietary level reduced after 23 days.
    d  Dietary level raised after 133 days.

    No toxicity could be demonstrated either by clinical appearance or by
    histopathological examination for isomers with 2 or 4 chlorine atoms
     ortho to the biphenyl bridge. The clinical signs observed in monkeys
    exposed to 3,4,3',4'-tetrachlorobiphenyl (up to 3 mg/kg diet) and
    3,4,5,3',4',5'-hexachlorobiphenyl (at 1 mg/kg diet) were similar in
    character and severity to those produced by Aroclor 1242 (at 100 mg/kg
    diet) and Aroclor 1248 (at 25 mg/kg diet) in monkeys (Allen et al.,
    1974a). The same histopathological lesions were found. The lesions of
    the skin and eyes were described as an expression of squamous atrophy
    or squamous cyst formation of the sebaceous glands. The epithelial
    changes in the skin and nails were found to be reversible. In this
    study, 2,5,2',5'-tetrachlorobiphenyl did not produce any clinical or
    pathological lesions at 3 mg/kg diet. Aroclor 1242 and 1248 were
    reported to contain about 0.24 and 0.34% of 3,4,3',4'-tetra-
    chlorobiphenyl, and it was concluded that this congener could account
    for some of the toxicity of the commercial mixtures (McNulty et al.,
    1980; McNulty, 1985).

    Rhesus monkeys exposed to 2,5,4'-trichlorobiphenyl showed a reversible
    primary injury of the arterioles, capillaries, and venules in the
    adrenal glands, kidneys, liver, brain, and lungs (Iatropoulos et al.,

    (b) Other animal species

    In studies on rats and mice, individual PCB-congeners caused adverse
    effects on the liver, spleen, and thymus. The most toxic compounds
    were the planar congeners 3,4,3',4'-tetrachlorobiphenyl,
    3,4,5,3',4'-pentachlorobiphenyl and 3,4,5,3',4',5'-hexachlorobiphenyl.

    Biocca et al. (1981) and Aulerich et al. (1985) compared the
    toxicities of various hexachlorobiphenyl isomers in mice. Male C57
    BL/6 mice were exposed via the diet to 0.3, 1, 3, 10, 30, 100, or
    300 mg 3,4,5,3',4',5'-hexachlorobiphenyl; 10, 30, 100, or 300 mg
    2,4,5,2',4',5'-hexachlorobiphenyl, 2,3,6,2',3',6'-hexachlorobiphenyl,
    or 2,4,6,2',4',6'-hexachlorobiphenyl/kg diet for 28 days. There were
    marked differences in dose response and in the severity of the
    pathological effects among the isomers. 3,4,5,3',4',5'-Hexa-
    chlorobiphenyl was the most toxic isomer causing mortality, and body
    and organ weight effects at all dose levels and was the only isomer
    that produced excess porphyrin accumulation. It was also the isomer
    that occurred in the highest concentration in the fat and the liver.
    3,4,5,3',4',5'-Hexachlorobiphenyl caused subcutaneous oedema,
    enlargement of the liver with accentuated hepatic lobular markings,
    fatty infiltration, hepatocellular swelling and necrosis, and atrophy
    of the thymus. The other 2 isomers caused the same lesions, but to a
    lesser extent.

    In the mice, the overall order of toxicity was 3,4,5,3',4',5'-
    hexa-chlorobiphenyl > 2,4,6,2',4',6'-hexachlorobiphenyl >
    2,4,5,2',4',5'-hexachlorobiphenyl > 2,3,6,2',3',6'-hexachlorobiphenyl,
    based on the effects on mortality and growth, and on histopathology.

    8.2.2  Intraperitoneal: reconstituted PCB mixtures

    Bandiera et al. (1984) administered reconstituted mixtures of PCDFs
    and reconstituted mixtures of PCBs, by intraperitoneal injection, to
    immature Wistar rats, to determine the effects on weight loss, thymic
    atrophy, and the induction of P-448 dependent monooxygenases. The
    mixtures consisted of compounds that persisted in the blood and liver
    of Yusho patients. From the results, it was clear that the PCDFs were
    600 to 2100 times more toxic than the PCBs.

    A group of 4, one-month-old, male Wistar rats received
    intraperitoneally a reconstituted PCB mixture containing 13 of the
    major congeners that have been identified in human milk, at the
    corresponding relative concentrations. The mixture was injected on
    days 1 and 3 in corn oil at dose-levels of 0, 0.45, 0.90, 4.5, or
    45 mg/kg body weight. The rats were killed on day 6 for
    histopathological studies. At the highest exposure level, increased
    relative liver weights and enlarged and vacuolated hepatocytes were
    observed together with changes in nuclei. In the thyroid, a mild
    reduction in follicular size, focal collapse, and changes in nuclei
    were found. No changes were seen with 0.9 mg/kg body weight (Gyorkos
    et al., 1985) (see section

    8.2.3   Dermal exposure

    (a) Aroclors

    Solutions of Phenoclor DP6, Clophen A60, or Aroclor 1260 in
    isopropanol were applied in doses of 118 mg on 50 cm2 of the shaved
    back skin of groups of 4 New Zealand rabbits, 5 times per week, for 38
    days. A group of 4 rabbits received the vehicle only. After initial
    reddening, transverse wrinkling and thickening of the skin developed
    with hyperplasia and hyperkeratosis of the epidermal and follicular
    epithelium. These effects were more marked with Clophen and Phenoclor
    than with Aroclors. During the study, deaths occurred in the Clophen-
    and Phenoclor-treated groups. Body weights and relative kidney weights
    were decreased in the Aroclor-treated rabbits. Histological liver
    changes were least marked in the Aroclor-treated rabbits. Treated
    rabbits had fluorescing livers and bone under UV radiation and also

    showed other evidence of porphyria. In the kidneys, hydropic
    degeneration of the convoluted tubuli, and tubular dilatation were
    found. There was atrophy of the thymic cortex and a reduction of
    germinal centres of the lymph nodes, as well as leukopenia, and some
    animals in all groups showed oedema of the abdominal and thoracic
    cavities, subcutaneous tissue, and pericardium. Faecal elimination of
    copro- and protoporphyrine was increased by all 3 PCBs, but was lowest
    with Aroclor 1260 (Vos & Beems, 1971).

    Puhvel et al. (1982) exposed groups of 3 female, hairless mice of 2
    strains, (Skh:HR-1 and HRS/J), topically to Aroclor 1254 (4 doses of 1
    or 8 mg/week, for 6 weeks) or Phenoclor 54 (5 doses of 0.2 mg/week,
    for 10 weeks) in acetone or an acetone-mineral oil-Tween 80 emulsion,
    or to the vehicle only. Punch biopsies of the skin were taken
    regularly. The skin of treated mice appeared grossly normal after the
    exposures, but examination of microscopic skin samples of
    Phenoclor-treated mice showed hyperkeratosis of the stratum corneum,
    epidermal hyperplasia, disappearance of sebaceous glands, and the
    presence of numerous keratinous cysts. No histological changes were
    observed in the internal organs.

    (b) Individual congeners

    In the study of Puhvel et al. (1982), described above, groups of 3
    hairless mice of both strains were also exposed topically to 5 doses
    of 0.2 mg 3,4,3',4'-tetrachlorobiphenyl/week, for 10 weeks. Grossly
    there were no visible changes. The histological changes induced in
    HRS/J mice were similar to those found after Phenoclor treatment.
    However, identical changes induced in Skh:HR-1 mice were already
    observed by 4 weeks. After 8 weeks, these lesions were more marked and
    also included hyperkeratosis of the sebaceous follicles and
    hyperkeratinization of intradermal pilar cysts. Often these cysts
    ruptured into the dermis leading to dense infiltrates of
    polymorphonuclear leukocytes. The treated mice showed weight gain,
    primarily because of large intra-abdominal fat deposits.

    In a dermal toxicity study on rabbits, with a protocol identical to
    that of the study of Vos & Beems (1971), skin lesions on animals
    treated with 2,4,5,2',4',5'-hexachlorobiphenyl were less severe than
    those on Aroclor 1260-treated animals. The liver damage observed
    histologically was essentially the same after treatment with either
    Aroclor 1260 or 2,4,5,2',4',5'-hexachlorobiphenyl, but the individual
    congener was more porphyrogenic (Vos & Nootenboom-Ram, 1972). In
    another study, 4 applications of a 25% solution of 3,4,3',4'-
    tetrachlorobiphenyl in olive oil to the inner surface of the ears
    of rabbits resulted in the same lesions that were found after 2
    applications of undiluted Kanechlor 400 or 500. The lesions included
    hyperkeratosis, dilatation of hair follicles, and the formation of
    keratinous cysts (Komatsu & Kikuchi, 1972).

    8.2.4   Appraisal

    Rhesus monkey is the most sensitive test species with regard to the
    general toxicity of PCBs, both as a mixture and as individual
    congeners. The toxicity of mixtures may be confounded by the presence
    of impurities, such as PCDFs, which are, or may have been, present in
    the mixtures tested. PCBs induce some of the biological and
    toxicological effects qualitatively similar to those induced by PCDFs.
    Another confounding variable in these studies is the difference in the
    composition of the mixtures (Aroclor 1242, 1248, 1254) used.

    Beating the above in mind, the available data show that Aroclor 1248,
    containing 4.4-8.7 ng of PCDFs/kg, still showed adverse general toxic
    effects in Rhesus monkeys at a dose of 0.1 mg/kg diet per day
    (0.09 mg/kg body weight per day) administered for an average of 18.2
    months (Bowman et al., 1981). A NOEL for general toxicity was not
    established for Aroclor 1248. The NOEL for the general toxicity of
    Aroclor 1242 was 0.04 mg/kg body weight per day, as established after
    dietary exposure for 133 days. The main effects induced by Aroclor
    1248 at 0.09 mg/kg body weight per day were an increased mortality
    rate, growth retardation, alopecia, acne, swelling of the Meibomian
    glands, and, possibly, immunotoxicity. Microscopically, enlarged
    hepatocytes, fatty liver with focal necrosis, and epithelial
    hyperplasia and keratinization of hair follicles were found. These
    effects appeared reversible. Aroclor 1254 at a dose level of
    0.200 mg/kg body weight per day showed several effects, not reported
    for 1248 (lymforeticular lesions, finger-nail detachment, gingival
    effects) and vice-versa (acne, alopecia), which could be related to
    the confounding variables noted above. Several effects observed in the
    monkeys exposed to Aroclor 1254 (hypertrophic gastropathy, bone marrow
    hyperplasia) were also observed in monkeys exposed to Aroclor 1248,
    but at a higher dose (4 mg/kg body weight per day). In contrast with
    the severe effects observed in adult Rhesus monkeys at low doses,
    relatively mild effects were shown by suckling Rhesus monkeys exposed
    to much higher doses.

    8.3  Skin and eye irritation, sensitization

    The injury to skin and eyelids following oral and/or dermal exposure
    to PCBs has been discussed in sections 8.2.1 and 8.2.3.

    8.4  Reproduction, embryotoxicity, and teratogenicity

    8.4.1  Reproduction and embryotoxicity  Oral

    (a) Mouse

    In castrated, mature, male NMRI mice (10-13 per group) that received
    28 daily doses of 0.25 mg Aroclor 1254/mouse, in peanut oil, the
    weight of the seminal vesicles was decreased, but this was not seen in
    intact mice (Orberg & Lundberg, 1974).

    When 23 adult female NMRI mice were mated with 22 control males and
    orally intubated with 0.025 mg Clophen A60/day, in peanut oil, for 62
    days prior to mating and up to days 8-10 of gestation, blastocytes
    failed to implant. Twenty-five (14 experimental and 11 control
    animals) out of the 45 animals were used to study the effects of PCBs
    on the estrous cycle. Effects, such as prolonged estrous cycle and
    less frequent periods of sexual receptivity and a decline in the
    number of implanted ova, were found (Orberg & Kihlstrm, 1973).

    In order to study the effects of PCBs on the development of sexual
    functions in the early postnatal period, these authors also mated mice
    that had been suckled by mothers dosed with Clophen A60 during the
    lactation period. A decrease in the frequency of implanted ova was
    noted, when both parents of the couple had been suckled with milk
    containing PCBs. When adult female NMRI mice received 50 mg of Clophen
    A60/kg body weight once per week, subcutaneously, during lactation,
    the same effect was observed in the offspring after mating with
    similarly exposed males (Kihlstrm et al., 1975).

    (b) Rat

    In a 2-generation reproduction study, groups of 10 male and 20 female
    Sherman weanling rats were fed diets containing 1, 5, 20, or 100 mg of
    Aroclor 1254/kg (equivalent to 0.06, 0.32, 1.5, and 7.6 mg/kg body
    weight, respectively), or diets containing 5, 20, or 100 mg/kg of
    Aroclor 1260 (equivalent to 0.39, 1.5, and 7.5 mg/kg body weight,
    respectively). Control groups comprised 20 male and 40 female rats.
    The F0 rats were started on the diet at 3-4 weeks of age and the F1
    rats, at weaning. The F0 rats were pair-mated when 3 and 7 months old
    to produce the F1a and F1b generations, respectively. Breeding-stock
    F1b rats were selected at weaning and pair-mated when 3 months old to
    produce the F2a, and, when 8 months old, the F2b generation. Rats
    exposed to Aroclor 1254 at 20 mg/kg diet or more showed a reduced
    litter size at birth, but not when exposed at weaning, in the F1b and

    F2 generations. At 100 mg/kg diet, the number of litters in the F2
    generation was decreased. In 2 F2a and 2 F2b litters no live offspring
    were found at birth, while pup survival at weaning was reduced in the
    F2a generation. At weaning, exposed F1a pups weighed less than their
    controls. Increased relative liver weights were found in male F1a
    weanlings at 1 mg Aroclor 1254/kg diet and in all weanlings at 5 mg/kg
    diet or more. Adult rats showed increased relative liver weights at
    levels of 20 mg/kg diet or more. No reproductive effects were found
    with 5 mg Aroclor 1254/kg diet. In groups treated with Aroclor 1260,
    increased liver weights were found in all weanlings at 5 mg/kg diet or
    more, but no effects on reproduction were seen, even at 100 mg/kg diet
    (Linder et al., 1974).

    The only effect observed in Holtzman rats fed Aroclor 1254 at a level
    of 500 mg/kg diet, for 3 weeks, was an increase in relative testes
    weights (Garthoff et al., 1977). Increased absolute testes weights and
    unchanged body weights were found in 6-month-old offspring of
    Sprague-Dawley dams exposed to daily doses of 30 mg Aroclor 1260/kg
    body weight in ethanol-sesame oil on days 14-20 of gestation. No
    effects on testes weight were found in animals treated with Aroclor
    1221 or Aroclor 1242 (Gellert & Wilson, 1979).

    Sager (1983) evaluated the effects on the reproductive function of
    adult male Holtzman rats following exposure of their mothers to doses
    of 8, 32, or 64 mg Aroclor 1254/kg body weight in peanut oil, on days
    1-3, 5, 7, or 9 of lactation. At all dose levels, 165-day-old males
    showed a decreased relative ventral prostate weight, and, at the 2
    higher doses, a decreased relative weight of seminal vesicles and
    testes as well as decreased body weight. In the ventral prostate,
    alveoli were decreased in number and showed folding of the mucosa and
    flattened epithelial cells. At 130 days of age, the male offspring at
    all 3 dose levels showed a reluctance to mate with control breeders,
    leading to a decreased number of pregnancies. Moreover, at the 2
    higher dose-levels, the females showed a reduced number of
    implantations and an increased resorption rate. The litters showed a
    reduced weight gain up to 11 days of age.

    This study was repeated with an evaluation of the reproductive
    performance of the second generation male rats from 130 days of age,
    following mating with normal females. In the first study, autopsy on
    pregnant females was carried out on day 11 or 12 of gestation. The
    females had fewer implants, fewer embryos, and a reduced proportion of
    ovulated eggs that implanted, compared with controls. The effects were
    dose-related. In a second study, females mated to the same males were
    autopsied on day 2 or 4 after mating. Sperm counts were not affected.
    At the 2 highest doses, fewer females had eggs in the expected state
    of development, the average number of blastocytes found in one uterine
    horn on day 4 was reduced, and an increased incidence of abnormally
    developed embryos was observed (Sager et al., 1987).

    Female Wistar rats exposed to daily doses of 10 mg Aroclor 1254/kg
    body weight, for at least one month, showed a prolonged estrous cycle,
    decrease in sexual receptivity, delay in timing of copulation, vaginal
    bleeding during gestation, decrease in litter size, and delay in the
    time of parturition. After mating of the rats to control males, the
    female offspring, exposed  in utero and during lactation, showed a
    slower rate of body weight gain, higher mortality, earlier vaginal
    opening, and a delay in the appearance of the first estrous cycle
    (Brezner et al., 1984).

    (c) Monkey

    Groups of 9 adult female Rhesus monkeys were exposed to a diet
    containing Aroclor 1248 (containing polychlorinated dibenzofurans) at
    levels of 2.5 or 5.0 mg/kg (Barsotti et al., 1976; Allen et al., 1979,
    1980). The study and the maternal toxicity data have already been
    described in section Within 4 months, menstrual bleeding and
    the duration of the menstrual cycle were increased. Flattening and
    prolongation of the serum progesterone peak during the menstrual cycle
    was observed. After 6 months of exposure, the females were mated with
    control males. Reproductive dysfunction was obvious as shown in Table
    44. Following the total exposure period of 18.2 months, the mothers
    were put on a control diet. Their menstrual cycles and serum
    progesterone levels returned to pre-exposure values. One year after
    exposure ceased, the females were again mated with control males and
    showed a recovery of their reproductive status (Table 44; Allen et
    al., 1980).

    Other groups of adult, female Rhesus monkeys were continuously fed
    diets with 0, 0.25, and 1.0 mg Aroclor 1016/kg (equivalent to 0, 0.01
    and 0.04 mg/kg body weight, respectively), in which no polychlorinated
    dibenzofurans were detected (no details). No maternal toxicity was
    noted at these levels. In this study, the females were mated with
    control males after 7 months of exposure. Reproductive dysfunction
    was not observed. Decreased birth weights were found in the offspring
    of mothers exposed to Aroclor 1016 at 1.0 mg/kg diet. Skin
    hyper-pigmentation occurred in both exposure groups (Barsotti & Van
    Miller, 1984). Preliminary reports have indicated possible effects on
    learning and behavioural tasks. The milk contained an average of 1.45
    and 3.92 mg/kg fat at 0.25 and 1.0 mg/kg, respectively, whereas the
    serum of the mothers contained 0.012 and 0.027 mg/litre, respectively
    (Heironimus et al., 1981; Levin & Bowman, 1983).

        Table 44.  Reproductive status of Rhesus monkeys exposed to dietary levels of Aroclor 1016 or 1248

    PCB          Exposure level             Total intake    Conceptions    Abortions       Stillborn     Live        Reference
    mixture                                   at                             and                        births
    (Aroclor)    Diet        Body           conception                     resorptions
                 (mg/kg)     weight         (mg/kg)

    1016         0           0                0              8/8           0/8             0/8            8/8        Barsotti &
                 0.25        0.01             8c             8/8           0/8             0/8            8/8        van Miller
                 1.0         0.04            30c             8/8           0/8             0/8            8/8        (1984)

    1248         0           0                0             12/12          0/12            0/12          12/12       Barsotti et al.
                 2.5         0.09a          105              8/8           3/8             0/8            5/8        (1976)
                 5.0         0.2            210              6/8           4/8             1/8            1/8

    1248         0           0                0              8/8           0/8             0/8            8/8        Allen et al.
                 2.5         0.09a          270              8/8b          1/8             0/8            7/8        (1980)
                 5.0         0.2            498              7/7b          1/7             2/7            4/7

    a  From: Bowman et al. (1981) amended.
    b  Breeding 1 year after exposure.
    c  Calculated assuming a body weight of 5 kg.

    In these studies, the monkeys were maintained on the diets during the
    gestation and lactation of the first generation. PCBs are known to
    cross the placenta and to be excreted via breast milk (see section
    6.3). The 6 infants born to monkeys during exposure at 2.5 or 5.0 mg
    Aroclor 1248/kg diet showed decreased birth weights, a small stature,
    and a decreased body weight gain during nursing. Within 2 months,
    signs of intoxication appeared including acne, increased skin
    pigmentation, swelling of the eyelids, and loss of eyelashes. In 3
    milk samples, values ranging from 0.154 to 0.397 mg PCBs/kg milk were
    measured, and, 1 milk sample contained 16.44 mg PCBs/kg fat. Three
    infants died. Necropsy and histopathology revealed atrophy of the
    thymus and lymph nodes, hypocellular bone marrow, moderate fatty
    infiltration of the liver, hypertrophic Meibomian glands, and
    keratinization of hair follicles. One dead infant showed hyperplasia
    of the gastric mucosa. The 3 surviving infants were weaned and
    subsequently showed marked improvements in their physical status
    (Allen & Barsotti, 1976; Allen et al., 1979, 1980). At 6 and 12 months
    of age, they were found to be hyperactive in a locomotor activity test
    and, between 7 and 24 months of age, they did not learn reversal tasks
    as readily as the controls. The PCB body burdens of these infants
    ranged between 11 and 27 mg/kg body fat, at the age of 8 months, and
    dropped to 0-1.6 mg/kg, at the age of 23 months. However, using the
    same apparatus, these infants showed hypolocomotor activity at 44
    months of age in comparison with the same controls (Bowman et al.,
    1978; Bowman & Heironimus, 1981; Bowman et al., 1982; Levin & Bowman,

    The infants delivered by the same adult females, after 1 year on a
    control diet, showed a reduced body weight and signs of intoxication
    similar to those observed in their siblings of the first breed.
    Two infants in each exposed group died. Milk samples contained
    0.02-0.19 mg PCBs/kg milk (Allen et al., 1980). At 12 months of age,
    when the PCB body burdens were only slightly higher than those of the
    controls, the infants showed hyperlocomotor activity (Bowman et al.,

    Groups of 4 or 6 Rhesus monkeys, which had been exposed to 0 or 2.5 mg
    Aroclor 1248/kg diet  in utero and during nursing until 4 months
    after birth, were tested at 4-6 years of age on delayed spatial
    alternation (DSA), a spatial learning and memory task. Deficits in
    performance accuracy were detected in 2 cohorts of monkeys, whose
    mothers had been fed 2.5 mg Aroclor 1248/kg diet for an 18-month
    period ending at least 12 months prior to pregnancy. The deficit was
    most apparent at the shorter delays, suggesting impairments in
    association or attention processes were involved rather than memory
    impairment. Such a deficit was also found in monkeys fed 1.0 mg
    Aroclor 1016/kg diet, but the effect was less pronounced. The
    appearance of a PCB-induced cognitive deficit more than 3 years after
    the end of exposure indicated the existence of long-term adverse
    consequences of perinatal PCB exposure (Levin et al., 1988).

    Clophen A30, which did not contain detectable levels of
    polychlorinated dibenzofurans (limit of detection <1 mg/kg), was
    given by gavage in a 1% solution of methylcellulose in water, once
    daily for 30 days, to 3 lactating Rhesus monkeys and their offspring
    at a level of 16 mg/kg body weight. PCB concentrations were measured
    in the serum of both mothers and infants and in the milk, on days -14,
    -7, 0, 1, 2, 4, 8, and then at weekly intervals until the end of the
    study. The mean PCB concentrations in the serum of mothers and infants
    were between 0.13 and 1.16 mg/litre and 0.07 and 2.67 mg/litre,
    respectively (before treatment days -14 and -7). The mean PCB levels
    in milk ranged from 0.63 mg/kg on day 1 of exposure to 18.90 mg/kg. On
    days -14 and -7, the concentrations in milk were 0.14 and 0.34 mg/kg
    (wet weight). Five control pairs were used. One dam and her offspring
    were sacrificed on day 22, exhibiting symptoms of anorexia,
    depression, lethargy, and ataxia. Two of 3 infants showed a decreased
    body weight gain. At autopsy of the infants, after the exposure
    period, slight degenerative changes were seen in the liver and the
    kidneys, together with slight demyelination of the central nervous
    system, slight to moderate gliosis of the cerebrum, and slight
    granular cell layer thinning of the cerebellum. On the basis of
    earlier work with adults in which the degenerative changes described
    above were considerable, the authors concluded that the nursing
    infants seemed to be less susceptible than the adults under the
    conditions of the studies (Iatropoulos et al., 1978; Bailey et al.,

    8.4.2   Teratogenicity  Aroclors (oral)

    (a) Mouse

    Haake et al. (1987) reported that treatment of pregnant C57Bl/6 mice
    with Aroclor 1254, by gavage, at 224 mg/kg body weight, on day 9 of
    gestation, did not result in any fetuses with cleft palate (see
    section 8.6.6).

    (b) Rat

    Wistar rats were exposed to dietary levels of Kanechlor 400 of up to
    250 mg/kg from day 1 to day 21 of gestation (Mizunoya et al., 1974).
    Fetuses showed decreased body weights from 10 mg/kg diet onward
    (equivalent to 0.67 mg/kg body weight), but did not show any increased
    incidence of malformations. Maternal toxicity was not observed and
    litter size and the number of litters were unaffected. Decreased pup
    survival was noted at dietary levels from 50 mg/kg (equivalent to
    3.5 mg/kg body weight) upwards. The 28-day-old offspring showed
    decreased body weight and increased relative liver weight from
    10 mg/kg.

    Commercial Kanechlor 300 or 500 was mixed with food and administered
    to pregnant Sprague-Dawley rats, throughout gestation, at levels of
    20, 100, or 500 mg/kg diet. On day 21, about three-quarters of the
    pregnant females were sacrificed; the remainder were allowed to litter
    naturally and the postnatal development of the pups was observed.

    Kanechlor 500 at a concentration of 500 mg/kg resulted in decreased
    maternal weight gain and decreased food consumption. At 20 and 500 mg
    Kanechlor 300/kg, and 500 mg Kanechlor 500/kg, the fetal weight
    decreased significantly. Resorption and malformations were not
    increased by treatment with Kanechlors. The Kanechlors did not show a
    teratogenic potential in this study (Shiota, 1976a).

    Offspring of Sprague-Dawley rats that received 20 mg of Kanechlor
    500/kg body weight on days 15-21 of gestation were slower than
    controls in achieving the water maze test at the age of 12-13 weeks,
    but did not perform worse in the open field test and in the swimming
    test (Shiota, 1976b). Behavioural effects were also observed in the
    offspring of ICR dams, 23-27 days of age, exposed to Aroclor 1254 in
    the diet at levels of 11 or 82 mg/kg (equivalent to 1.7 and 12 mg/kg
    body weight) from 3 days before mating up to weaning. The offspring
    were maintained on the same diet. PCB exposure did not have any
    effects on the ability to learn an avoidance response, but increased
    the latency to make such a response. Moreover, the young mice
    exhibited slower habituation to an open field (Storm et al., 1981).

    The effects of Fenchlor 42 (trichloro- 63%, tetrachloro- 33%, and
    small amounts of dichlorobiphenyl; purity 97.5%) exposure of Fischer
    344 male and female rats were studied through assessment of the
    behavioural development of their F1 progeny. Female rats were
    administered 5 daily ip injections of corn oil or 5-10 mg Fenchlor
    42/kg body weight per day, 2 weeks prior to mating. Another group
    received 2-4 mg/kg per day during gestation (days 6-15 of pregnancy)
    and a third group of 8 previously treated lactating females received
    corn oil or 1-2 mg/kg per day on postnatal days 1-21. The total doses
    in the 3 groups were 25-50, 20-40 and 20-40 mg/kg body weight.
    Dose-dependent differences in behaviour were found in the offspring of
    the PCB-treated animals. Differences in the development of cliff
    avoidance reflexive behaviour, swimming ability, and open field
    activity were evident. The PCB exposure of female animals during
    gestation and lactation resulted in impaired acquisition of the active
    avoidance behaviour, while preconception PCB exposure affected active
    avoidance performance, as reflected in an increased number of
    avoidance responses to reach criterion for extinction (Pantaleoni et
    al., 1988).

    Doses of 0, 6.25, 12.5, 25, 50, or 100 mg/kg body weight per day of
    Aroclor 1254 were administered to rats, by gavage, on days 6-15 of
    gestation. Average pup weights were reduced at 100 mg/kg, though total
    litter weight (average weight times number of fetuses) did not differ
    from controls. There were no skeletal or visceral abnormalities or
    effects on conception, resorptions, litter size or number, or average
    litter weight in any of the treated groups (Villeneuve et al., 1971b).

    Spencer (1982) exposed Holtzman rats to a diet containing Aroclor 1254
    at levels of 25 up to 900 mg/kg diet from day 6 to day 15 of gestation
    and found reduced maternal body weight gain and decreased fetal
    survival at birth from 300 mg/kg diet (equivalent to 18 mg/kg body
    weight) upwards, and decreased fetal weights at birth from 100 mg/kg
    (equivalent to 8 mg/kg body weight) upwards. No visceral or skeletal
    data were available.

    When Sherman rats were exposed to doses of Aroclor 1254 of up to
    100 mg/kg body weight, in peanut oil, from day 7 to day 15 of
    gestation, a decrease in the survival of the pups was found. At
    weaning, the survival and body weight of the grossly normal pups were
    reduced at the 100 mg/kg dose level, but not at 50 mg/kg body weight.
    No effects were observed at 100 mg Aroclor 1260/kg body weight (Linder
    et al., 1974).

    (c) Monkey

    Two pregnant Cynomolgus monkeys  (Macaca fascicularis) were dosed
    with Aroclor 1254 at 100 mg/kg body weight per day and 1 monkey, at
    400 mg/kg body weight per day, from day 60 of gestation. One control
    animal received the vehicle, corn oil. The 2 monkeys fed 100 mg/kg
    delivered dead male infants after 105 and 108 days of dosing, and the
    female fed 400 mg/kg delivered a female infant (with no overt clinical
    signs of toxicity) that died at 139 days of age with an acute
    bronchopneumonia. The breast milk of the monkey fed 400 mg/kg
    contained, over a period of 5-75 days after parturition,
    concentrations of 73.7 up to 139.4 mg/kg, on a fat basis. No overt
    signs of toxicity were observed in the adult animals, with exception
    of finger nail loss. All 3 treated monkeys showed impaired
    immunological capacity, assessed at approximately 50 days postpartum
    (148 days of treatment) (Truelove et al., 1982).

    (d) Rabbit

    Rabbits were exposed to 0, 1.0, or 10.0 mg Aroclor 1254/kg body weight
    and in another study to 0, 12.5, 25.0, or 50 mg/kg body weight (purity
    not reported), by gavage, on days 1-28 of pregnancy. Abortions, still
    births, and maternal deaths occurred at 12.5 mg/kg body weight or
    more, but no teratogenic effects were found (Villeneuve et al.,
    1971a,b).  Aroclors (subcutaneous)

    (a) Mouse

    A possible teratogenic effect was observed in ddy mice subcutaneously
    exposed to doses of 1-5 mg Kanechlor 500/mouse (equivalent to
    40-200 mg/kg body weight) in 95% ethanol from day 6 to day 15 of
    gestation. A dose-related increase in maternal mortality was observed
    from a dose of 3 mg/mouse onwards. Some dams showed skin lesions,
    alopecia, or swelling of the liver, but no effects on body weight
    gain. A slight increase was noted in the incidence of dead and
    resorbed fetuses. The incidence of cleft palate was increased in a
    dose-related manner from the lowest dose (Watanabe & Sugahara, 1981).  Individual congeners (oral)

    (a) Mouse

    Orberg (1978) fed groups of 20-56 pregnant female NMRI-mice 0, 0.05,
    or 0.5 mg 2,5,4'-trichloro- or 2,4,5,2',4',5'-hexachlorobiphenyl,
    dissolved in peanut oil, per animal, from days 1 to 6 of gestation. A
    significant decrease in the pregnancy of implanted ova was found with
    the 0.5 mg treatment. No effects on percentage of pregnancies and mean
    number of corpora lutea were found. There were no effects at the lower
    dose level.

    A dose-related increase in embryotoxicity and the incidence of
    malformed fetuses, mainly showing cleft palate and hydronephrosis, was
    found in pregnant CD-1 mice after exposure to daily doses of 2, 4, 8,
    or 16 mg 3,4,5,3',4',5'-hexachlorobiphenyl/kg body weight, in
    cotton-seed oil, on days 6-15 of gestation. Lower dose levels, e.g.,
    0.1 and 1 mg/kg body weight were without effects. No dibenzofurans
    were detectable (no details) in the test compound. The dams showed a
    decreased body weight gain at 8 mg/kg body weight. The authors
    reported that 3,4,3',4'-tetrachlorobiphenyl and 2,3,4,2',3',4'-hexa-
    chlorobiphenyl also produced the same teratogenic effects, though they
    were less potent than those of 3,4,5,3',4',5'-hexachlorobiphenyl
    (Marks et al., 1981).

    A neurobehavioural "spinning" syndrome (a syndrome characterized by
    the fact that the mice rotate or spin in a circular motion when held
    by the tail) and hydronephrosis developed in CD-1 mouse weanlings,
    whose dams received, by gavage, 32 mg 3,4,3',4'-tetrachlorobiphenyl/kg
    body weight, in corn oil, on days 10-16 of gestation. Maternal
    neurotoxicity was not observed. Histological and ultrastructural
    examination of the CNS of affected mice revealed longitudinal
    projections of the cylindrical CNS in the ventral and dorsal roots
    and, to a lesser extent, in cranial nerve roots. The effect was
    possibly related to an observed altered development of striatal
    synapses (Chou et al., 1979; Tilson et al., 1979; Agrawal et al.,

    (b) Rat

    The congener 3,4,3',4'-tetrachlorobiphenyl was found to be embyrotoxic
    and caused accumulation of blood in the amniotic fluid and the
    gastrointestinal tract of fetuses from Sprague-Dawley rats treated on
    days 6-18 of gestation with doses of 3 or 10 mg/kg body weight in corn
    oil. Decreased fetal growth was also observed (Wardell et al., 1982).
    When the rats were allowed to deliver, high perinatal mortality was
    observed, which appeared to be related to an increase in gestational
    length and to be independent of the smaller total litter size. In
    addition, pup weights were found to be lower than those of controls
    (Rands et al., 1982; White et al., 1983).

    (c) Guinea-pig, mouse

    Neither embryotoxicity nor teratogenicity were found in Dunkin Hartley
    guinea-pigs, and CBA mice exposed  in utero to 2,4,5,2',4',5'-
    hexa-chlorobiphenyl during gestation (Mattsson et al., 1981; Brunstrm
    et al., 1982; Aulerich et al., 1985) and in C57BL/6N mice similarly
    exposed to 2,4,5,2',4',5'- or 2,3,4,5,3',4'-hexachlorobiphenyl
    (Birnbaum et al., 1985).

    Pregnant guinea-pigs received a total dose of 100 mg technical grade
    Clophen A50 orally, in peanut oil, from day 16 to day 60 of gestation,
    25 mg 2,4,5,2',4',5'-hexachlorobiphenyl from day 16 to day 60 of
    gestation, or 100 mg from day 22 to day 60 of gestation. The
    administration of Clophen A50 resulted in fetal deaths, but no
    maternal deaths. In contrast, 2,4,5,2',4',5'-hexachlorobiphenyl did
    not cause fetal deaths. Prenatal weight of live fetuses was increased
    by a dose of 25 mg, but not by 100 mg of 2,4,5,2',4',5'-hexa-
    chlorobiphenyl (Brunstrm et al., 1982).

    (d) Monkey

    In a briefly reported study, 6 female Rhesus monkeys received 9 doses
    of 70 or 350 g 3,4,3',4'-tetrachlorobiphenyl/kg body weight by
    gavage, in corn oil, from day 20 to day 40 of gestation. A control
    group comprised 12 animals. Maternal toxicity (not specified) but no
    mortality, was reported in all exposed monkeys from day 31 following
    exposure. Between days 17 and 35 following exposure all exposed
    fetuses and 3 out of 12 control fetuses aborted (McNulty, 1985).

    8.4.3  Appraisal

    The Rhesus monkey is the most sensitive species with regard to general
    toxicity (see section 8.2) and particularly with regard to
    reproductive toxicity. The presence of PCDFs and the variation in PCB
    composition may be confounding factors in determining the reproductive
    toxicity of PCB mixtures. Aroclor 1248, containing 4.4-8.7 g
    PCDFs/kg, adversely affected the reproductive performance of female

    Rhesus monkeys, mated with control males after 6 months of dietary
    exposure to a toxic dose of 0.09 mg/kg body weight per day and
    continuation of the exposure for an average of up to 10 months. This
    effect was reversible after an exposure-free period of 1 year. No
    effect on reproduction was found in female Rhesus monkeys exposed to a
    non-toxic dose of 0.01 or 0.03 mg Aroclor 1016 (reported not to
    contain PCDFs)/kg body weight per day and mated after 7 months with
    control males.

    Neonates of the nursing mothers exposed to Aroclor 1248 showed adverse
    effects similar to those seen in their mothers and, in addition,
    persistent behavioural disturbances, atrophy of the thymus and lymph
    nodes, bone marrow hypoplasia, and hyperplasia of the gastric mucosa.
    Neonates of the mothers after the recovery period still showed adverse
    effects, as well as the neonates of the mothers exposed to Aroclor
    1016. These effects were caused by PCBs, with or claimed to be
    without, PCDFs, transmitted via the placenta during gestation and
    later via the breast milk. Neonates have much greater susceptibility
    to PCB toxicity when exposed via the mothers compared with suckling
    monkeys orally exposed. Reproductive toxicity has also been observed
    in the mink, rabbit, and rat. The changes seemed to be related to
    alterations in the serum levels of gonadal steroid hormones, as a
    result of enzyme induction. PCBs may also bind to the cytoplasmic
    estrogen receptor. Effects have also been observed on the estrus cycle
    of female rats, minks, and monkeys, on the sex organs of male rats,
    and on the implantation rate of fertilized ova following exposure of
    female mice or male rats.

    Comprehensive teratological examinations have not been conducted;
    however, the available studies indicated that the Aroclors were not
    teratogenic in rats and nonhuman primates, when tested via the oral
    route during the critical periods of organogenesis at doses that
    produced fetotoxicity and/or maternal toxicity. Although the
    fetotoxicity of Aroclors is documented in several species of animals,
    the possibility that contaminants (e.g., PCDFs) might be (partly)
    responsible for the effects should be recognized.

    The results of the reproduction and teratogenicity studies are
    summarized in Tables 45, 46, and 47.

    8.4.4  Mutagenicity and related end-points  DNA damage

    PCBs have been shown to interact with the proteins, RNA and DNA, after
    metabolic activation. The potential of readily metabolizable
    PCB-congeners to cause primary DNA damage was indicated by the
    activity of 2,5,2',5'-tetrachlorobiphenyl and its metabolites in
    causing DNA, single-strand breaks in an alkaline elution assay with

    L-929 cells  in vitro (Stadnicki et al., 1979). Furthermore,
    unscheduled DNA synthesis was elicited by 4-chlorobiphenyl  in vitro
    in Chinese hamster ovary cells (Wong et al., 1979). No unscheduled DNA
    synthesis was elicited by Aroclor 1254 in rat hepatocytes  in vitro
    (Probst et al., 1981).

    DNA-breaking activity was found in an alkaline elution assay with
    hepatocytes of intact rats treated  in vivo with a single, high dose
    (500 mg Aroclor 1254/kg body weight, intraperitoneally, or 1295 mg/kg
    body weight, orally) with complete repair of the damage within 48 h
    (Robbiano & Pino, 1981). Aroclor 1254 was also shown to be a
    DNA-breaking agent in an alkaline elution assay  in vitro with rat
    hepatocytes (Sina et al., 1983). An alkaline sedimentation assay
    showed the DNA-breaking activity of Aroclor 1254 in rats treated
     in vivo with a single intraperitoneal dose of 500 mg/kg body weight.
    In this assay, the same Aroclor 1254 pretreatment of the rats  in vivo
    elevated the DNA-breaking activity of the direct-acting, alkylating
     N-methyl- N'-nitro- N-nitrosoguanidine and the carcinogens,
    dimethylnitrosamine and benzo (a)pyrene,  in vitro (Mendoza-Figueroa
    et al., 1985).  Mutagenicity tests

    Many mutagenicity tests have been carried out over the years, with
    different PCB mixtures. Most of these were commercial mixtures the
    composition of which was not described. Only a few studies are
    available on specific congeners. Besides studies on microorganisms,
    mammalian cell point mutation, dominant lethal assays, micronucleus
    tests, chromosome and cytogenicity studies, and DNA repair studies
    were carried out. With a few exceptions the results of most of the
    studies with PCB mixtures were negative (see Table 48).

    Aroclor mixtures and the congener 2,5,2',5'-tetrachlorobiphenyl and
    its metabolites did not induce point mutations in  Salmonella
     typhimurium TA 98, TA 100, TA 1535, TA 1537, and TA 1538 with, and
    without, metabolic activation (Wyndham et al., 1976; Hsia et al.,
    1978; Bruce & Heddle, 1979; Schoeny et al., 1979; Shahin et al.,
    1979), neither did 2,4,2',4'- and 3,4,3',4'-tetrachlorobiphenyl and
    2,4,6,2',4',6'-hexachlorobiphenyl in the strains TA 98 and TA 100
    (Schoeny, 1982).

    However, Wyndham et al. (1976) found a dose-related mutagenic activity
    of 4-chlorobiphenyl and, to a lesser extent, of Aroclor 1221 in the
    strain TA 1538, after metabolic activation by the S9 liver fraction of
    uninduced rabbits. The study was repeated 3 times in the same
    laboratory, but the results of Wyndham et al. (1976) could not be
    confirmed (Safe, 1980). Schoeny (1982) could not detect any mutagenic
    activity of 4-chlorobiphenyl in the same dose range in the strains TA
    98, TA 100, TA 1535, and TA 1537, with, or without, the S9 liver
    fraction of induced rats.

        Table 45.  PCBs: reproduction and embryo toxicity

    Animal (strain,  PCB mixture    Exposure period    NOAEL          LOAEL          Parameters, effects               Reference
    sex)                            (oral)             (mg/kg body    (mg/kg body
                                                       weight)        weight)

    Rat (Sherman,    Aroclor 1254   continuous up to                  0.06 (1254)    increased relative liver          Linder et al.
    10 male,         Aroclor 1260   weaning                                          weights in male F1A               (1974)
    20 female)                                                                       weanlings
                                                                      0.32 (1254)
                                                                                     former effect in all weanlings
                                                       0.32 (1254)
                                                       7.5 (1260)                    reproduction parameters
                                                                                     reproduction parameters

    Rhesus monkey    Aroclor 1016   7 months           0.03 (0.04)                   reproduction parameters           Barsotti &
    (female)                                                                         [0.03(0.04): decreased birth      van Miller
                                                                                     weight] [0.01: skin               (1984)

    Rhesus monkey    Aroclor 1248   6 months                          0.09           abortions, resorptions, live      Barsotti et al.
    (female)                                                                         births                            (1976)
                                    1 year after       0.09           0.2                                              Allen et al.
                                    exposure                                         stillborn, live births            (1980)

    Table 46.  PCBs: teratogenicity

    Animal (strain)    PCB-mixture     Exposure period    NOAEL         LOAEL           Parameters, effects         Reference
                       (oral)                             (mg/kg body   (mg/kg body
                                                          weight)       weight)

    Rat                Aroclor 1254    days 6 to 15 of     50            100            reduced average pup         Villeneuve et al.
                                       gestation                                        weight                      (1971b)

    Rat (Holtzman)     Aroclor 1254    days 6 to 15 of    < 8              8            decreased fetal weight at   Spencer (1982)
                                       gestation                                        birth

    Rat (Sherman)      Aroclor 1254    days 7 to 15 of    100                           reproduction effects        Spencer (1982)

    Rabbit             Aroclor 1254    days 1 to 28 of     10             12.5          abortions, still births,    Villeneuve et al.
                                       gestation                                        maternal deaths             (1971 a,b)

                       Aroclor 1221    days 1 to 28 of     25                           fetotoxicity

    Table 47.  PCBs: teratogenicity of individual congeners

    Animal         PCB-mixture      Exposure period     NOAEL           LOAEL           Parameters, effects                  Reference
    (strain)                        (oral)              (mg/kg body     (mg/kg body
                                                        weight)         weight)

    Mice           2,5,4'-TCB       day 1 to 6 of       0.05/animal     0.5/animal      decrease in the number               rberg (1978)
    (NMRI)         2,4,5,2',4',     gestation                                           of implant/dams

    Mice           3,4,5,3',4',     day 6 to 15 of                      2               embryotoxicity, malformations        Marks et al.
    (CD-1)         5'-HCB           gestation                                           (cleft palate, hydronephrosis)       (1981)

                   3,4,3',4'-TCB                                        8               maternal toxicity, less potent
                   2,3,4,2',                                                            than that of 3,4,5,3',4',5'-HCB

    Rat            3,4,3',4'-TCB    day 6 to 18 of                      3               embryotoxicity                       Wardell et al.
    (Sprague-                       gestation                                                                                (1982)

    Rhesus         3,4,3',4'-TCB    day 20 to 40 of                     0.07            maternal toxicity, total abortions   McNulty (1985)
    monkey                          gestation

    Table 48.  Results of mutagenicity, and related, tests

    Chemical         Test system        Strain        Dose                  Metabolic            Result                Reference
    substance                                                               activation

    2,5,2',5'-       Salmonella         TA 1538       200 g/plate          modified             -                     Wyndham et al.
    tetrachloro      typhimurium                      100 g/plate          microsomal           -                     (1976)
    biphenyl                                          50 g/plate           fraction             -
                                                      10 g/plate           from rabbits         -

    Aroclor 1268     Salmonella         TA 1538       200 g/plate          modified             -                     Wyndham et al.
                     typhimurium                      100 g/plate          microsomal           -                     (1976)
                                                      50 g/plate           fraction             -
                                                      10 g/plate           from rabbits         -

    Aroclor 1254     Salmonella         TA 1535       4 different           both with, and       -                     Heddle & Bruce
                     typhimurium        TA 1537       concentrations,       without, S-9 in      -                     (1977)
                                        TA 98         figures not           all cases            -
                                        TA 100        given                                      -

    Aroclor 1254     Salmonella         TA 1535       8 different           both with, and       -                     Schoeny et al.
                     typhimurium        TA 1537       concentrations        without, S-9         -                     (1979)
                                        TA 98         from 0.5 to                                -
                                        TA 100        500 g/plate                               -

    Aroclor 1254     Salmonella         TA 1535       0.05, 0.5,            both with, and       -                     Bruce & Heddle
                     typhimurium        TA 1537       5, 50, and            without, S-9         -                     (1979)
                                        TA 98         500 g/plate                               -
                                        TA 100                                                   -

    Table 48.  (cont'd).

    Chemical         Test system        Strain        Dose                  Metabolic            Result                Reference
    substance                                                               activation

    Aroclor 1254     Salmonella         TA 1538       50, 100, 500,         both with, and       -                     Shahin et al.
                     typhimurium        TA 98         1000, 2000,           without, S-9         -                     (1979)
                                                      5000 g/plate

    Aroclor 1242     V79                              50, 100, and          without              -                     Hattula (1985)
    Clophen A60      hamster cells                    150 g/ml             metabolizing         -
                     co-cultivated                                          cells
                     with lethally

    Aroclor 1254     Micronucleus                     4 different                                -                     Heddle & Bruce
                     test                             concentrations,                                                  (1977)
                     (erythrocytes)                   figures not

    Aroclor 1254     Micronucleus       (C57B1/6     (approximately)                            -                     Bruce & Heddle
                     test               C3H/He)       LD50, 1/2, 1/4,                            (at all doses)        (1979)
                                        F1 mice       and 1/8 of the
                                                      highest dose
                                                      (5 consecutive
                                                      days, ip)

    Table 48.  (cont'd).

    Chemical         Test system        Strain        Dose                  Metabolic            Result                Reference
    substance                                                               activation

    Kanechlor 500    Micronucleus       ddY-mice      100 mg/kg in          -                    -                     Watanabe et al.
                     test with                        corn oil orally                                                  (1982)
                     polychromatic                    and 100 mg/kg
                     erythrocytes                     95% ethanol

    Aroclor 1254     Chromosomal        (embryonic    0 or 10 mg/kg                              inconclusive          Peakall et al.
                     aberrations        Ring          diet                                                             (1972)

    Aroclor 1254     Chromosomal        (human        100 mg/litre                               -                     Hoopingarner
                     aberrations        lymphocytes)  culture medium                                                   et al. (1972)

    Aroclor 1254     Chromosomal        male          5, 50, 500 mg/kg                           negative at all       Garthoff et al.
                     abnormalities in   Holtzmann     diet                                       doses                 (1977)
                     bone marrow and    rats

    Aroclor 1242     Chromosomal        Osborne-      5000 mg/kg  1a                            -                     Green et al.
                     abnormalities      Mendel        2500 mg/kg  1                             -                     (1975a)
                     in bone marrow     rats          1250 mg/kg  1                             -
                     cells and          500 mg/kg  4                       -

    Table 48.  (cont'd).

    Chemical         Test system        Strain        Dose                  Metabolic            Result                Reference
    substance                                                               activation

    Clophen A30      Drosophila                       62.5, 125, 250,                            -b                    Nilsson & Ramel
                     melanogaster                     and 500 mg/litre                                                 (1974)
                     (adults or larvae) substrate

    Clophen A50      Drosophila                       25, 50, 100,                               -b                    Nilsson & Ramel
                     melanogaster                     and 200 mg/litre                                                 (1974)
                     (adults or larvae) substrate

    Aroclor 1242     Dominant Lethal    Osborne-      2500 mg/kg  1a                            -                     Green et al.
                     test               Mendel        1250 mg/kg  1                             -                     (1975b)
                                        rats          625 mg/kg  1                              -
                                                      250 mg/kg  5                              -
                                                      125 mg/kg  5                              -

    Aroclor 1254     Dominant Lethal    Osborne-      150 mg/kg  5a                             -                     Green et al.
                     test               Mendel        75 mg/kg  5                               -                     (1975b)

    Aroclor 1254     Dominant Lethal    Osborne-      25, 100 mg/kg                              -                     Green et al.
                     test               Mendel        diet for 70 days                           -                     (1975b)

    Table 48.  (cont'd).

    Chemical         Test system        Strain        Dose                  Metabolic            Result                Reference
    substance                                                               activation

    Aroclor 1254     Sperm              (C57 B1/6    (approximately)LD50,                       negative at all       Bruce & Heddle
                     Abnormality        C3H/He)F1     1/2, 1/4, 1/8 top dose                     doses                 (1979)
                                        mice          on 5 consecutive
                                                      days, ip

    Aroclor 1254     mitotic index      human         100 mg/litre                               mitotic index         Hoopingarner
                                        lymphocytes   culture medium                             equivocal             et al. (1972)

    4-chloro-        DNA repair and     Chinese       10-5 mmol/litre                            covalent-             Wong et al.
    biphenyl and     unscheduled        hamster       3H-4-chloro-                               binding to            (1979)
    metabolites      synthesis          ovary cells   biphenyl, 24 h                             protein, RNA,
                     (hydroxyurea                                                                and DNA
                     addition suppress                                                           Increase
                     DNA synthesis)                                                              specific
                                                                                                 activity with

    a  Means single dose ( 1) or 5 doses in 5 days ( 5).
    b  No loss or non-disjunction of sex-chromosomes.
    SD=Significant decrease.

    Dose-related chromosome breakage was found in human lymphocytes
    exposed to the planar PCB congener, 3,4,3',4'-tetrachlorobiphenyl, at
    0.1-10-4 g/ml (Sargent et al. (1989). In contrast, the non-planar
    2,5,2',5'-tetrachlorobiphenyl did not cause chromosome damage in a
    comparable test, even at concentrations as high as 1 g/ml. However, a
    combination of 3,4,3',4'-tetrachlorobiphenyl at a concentration of
    10-5 g/ml with 2,5,2',5'-tetrachlorobiphenyl caused chromosomal
    damage that was far in excess of what might be expected from higher
    doses of 3,4,3',4'-tetrachlorobiphenyl alone. The results suggest that
    some PCB congeners may interact to cause synergistic effects.

    Peakall et al. (1972) carried out cytogenic studies on Ring dove
    embryos  (Streptopelia risoria); 6 embryos were from dove pairs not
    fed PCBs (controls) and 17 embryos were from PCB-fed (10 mg/kg diet)
    pairs. The frequencies of chromosome aberrations were recorded for
    chromosome pairs occurring in metaphase cells of allantoic sac and
    limb bud origin. Mean aberration rates were as follows: control 0.8%
    (range 0-2.0%) and PCB-treated 1.8% (range 0-9.4%). It was concluded
    by the authors that these results were indicative of a possible
    clastogenic action of PCBs.

    A DNA repair assay was carried out by Wong et al. (1979) using CHO
    cells and measuring the effects of 4-chlorobiphenyl (10 mol/litre) on
    the unscheduled DNA synthesis (UDS) in the presence of hydroxyurea
    (HU), a chemical agent that suppresses normal replicative DNA
    synthesis. The quantification of DNA synthesis was determined by the
    uptake of [H3]-thymidine into the cellular DNA. A significant
    (1.6-fold) enhancement of UDS was found when the cells were incubated
    for 2.5 h in the presence of HU, 4-chlorobiphenyl, and thymidine.  Cell transformation

    Aroclor 1254 did not cause an increase in benzo (a)pyrene-induced
    transformation in a test using C3H10 T1/2 CL8 mouse embryo fibroblasts
    (Nesnow et al., 1981).

    Aroclor 1254 also failed to transform Golden Syrian hamster cells
    76-582 in culture at 50 l/ml (Pienta, 1980).  Cell to cell communication

    The congener 2,4,5,2',4',5'-hexachlorobiphenyl inhibited one form of
    intercellular communication in V79 Chinese hamster cells, i.e.,
    metabolic cooperation by mutant rescue at non-cytotoxic levels, while
    3,4,5,3',4',5'-hexachlorobiphenyl was inactive (Tsushimoto et al.,
    1983).  Interaction

    Grolier et al. (1989) studied the effects of Vitamin A dietary intake
    (2 and 20 IU/g of food) on the mutagenicity of benzo (a)pyrene
    (B (a)P) towards  Salmonella typhymurium TA 98, either in control
    rats or in animals treated with 2,4,5,2',4',5'- hexachlorobiphenyl and
    3,4,3',4'-tetrachlorobiphenyl. The planar tetrachlorobiphenyl strongly
    increased B (a)P-monooxygenase activity and glutathione transferase,
    while the non planar hexachlorobiphenyl was a strong inducer of
    epoxide hydrolase and a weak inducer of B (a)P-monooxygenase. Enzyme
    induction was not modified by changes in Vitamin A intake. A greater
    mutagenic effect was observed in the tetrachlorobiphenyl group than in
    the hexachlorobiphenyl group. This could be related to the specific
    form of cytochrome P-450 induced by the tetrachlorobiphenyl congener.
    In PCB-treated rats, the mutagenic activity of B (a)P was higher in
    the 20-IU group than in the 2-IU group.  Cell division parameters

    Tests on Osborne-Mendel rats gave various results, but may provide the
    most important clue to the mechanism of action of PCBs in
    carcinogenesis. At high doses (5000 mg Aroclor 1242/kg given once, and
    500 mg/kg, given in a series of 4 daily doses), there were significant
    decreases in the numbers of spermatogonial cells in mitosis. Single
    dose levels of 1250 or 2500 mg/kg gave negative results (Green et al.,
    1975a). Garthoff et al. (1977) also found negative results in male
    Holtzmann rats treated with 0, 5, 50, or 500 mg Aroclor 1254/kg diet
    for 5 weeks with regard to the mitotic indices of bone marrow and
    spermatogonial cells. The data of Hoopingarner et al. (1972) showed an
    increase in mitotic index in human lymphocytes exposed to Aroclor

    The above-mentioned studies provide evidence that Aroclors can enhance
    cell proliferation, and this is of special interest because it
    suggests that the Aroclors may act to promote carcinogenesis to a
    greater extent than to initiate it. The effect on cell proliferation
    requires further examination in a variety of systems.

    8.5  Carcinogenicity

    Hayes (1987) critically reviewed the available evidence for, and
    against, the view that environmental PCBs present a significant
    potential carcinogenic hazard for humans.

    8.5.1  Long-term toxicity/carcinogenicity

    (a) Mouse

    Nagasaki et al. (1974) exposed 10 groups of dd mice (6-12 of each sex)
    to Kanechlor 300, 400, or 500 in the diet at levels of 0, 100, 250, or
    500 mg/kg (equivalent to 0, 5, 12.5, and 25 mg/kg body weight,
    respectively) for 32 weeks. Nine nodular hyperplasia and 7
    hepatocellular carcinomas were found in 17 male mice, and 4 cases of
    liver hypertrophy in 17 female mice, after exposure to 500 mg
    Kanechlor 500/kg diet. No neoplasms were found in the other groups.

    In another mouse study, groups of male BALB/cJ mice were fed 0 or
    300 mg Aroclor 1254/kg diet (equivalent to 50 mg/kg body weight) for 6
    or 11 months. Adenofibrosis was observed in the livers of all 22 mice
    fed Aroclor 1254 for 11 months, but not in those of the 24 mice
    exposed for 6 months. Hepatomas were noted in 9/22 mice exposed for 11
    months and in 1/24 mice exposed for 6 months. No tumours were found in
    the controls (Kimbrough & Linder, 1974).

    (b) Rat

    In a preliminary study, liver tumours (multiple adenomatous nodules)
    were induced by Kanechlor 400 (containing 2,4,3',4'-; 2,5,3,3'-;
    2,3,4,4'-; and 3,4,3',4'-tetrachlorobiphenyls) in 6/10 females, but
    not in male Donryu rats, in 400 days. The dietary exposure was
    periodically adjusted according to animal weights and ranged from 38.5
    to 616 mg/kg diet. The latter dose level was administered for 275
    days. The number of animals used was small (10 treated and 5 control
    rats of each sex). Increased incidences of pneumonia, and lung and
    intracranial abscesses were found in rats on a diet containing
    Kanechlor 400, possibly because of lowered resistance to infection
    (Kimura & Baba, 1973).

    Ito et al. (1974) fed 10 groups of 29 male Wistar rats with Kanechlor
    300, 400, or 500 at dose levels of 100, 500, or 1000 mg/kg diet
    (equivalent to 5, 25, and 50 mg/kg body weight, respectively) for
    28-52 weeks. Another group received the control diet. A number of
    animals died in all groups (4 up to 21 animals); within the treated
    groups, deaths were more or less dose related. Adenofibrosis was
    observed in the livers of rats fed 1000 mg/kg diet of all 3 mixtures.
    Kanechlor 500 produced nodular hyperplasia at all dose-levels and at a
    higher incidence than the mixtures with a lower chlorine content. No
    neoplastic nodules were observed in the controls. Kanechlor 300 and
    500 (100 mg/kg diet) did not show significant growth inhibition or
    increases in liver weight. No fibrosis or cholangiofibrosis, bile duct
    proliferations, fatty changes, or cellular hypertrophy in the liver
    were found. The liver nodular hyperplasia, designated as preneoplastic
    by the investigators, was found in 3/25 of the Kanechlor 500

    (100 mg/kg diet) treated animals, 1/22 of the Kanechlor 300 (100 mg/kg
    diet) treated animals, and 0/18 controls. The 1/22 (4.5% incidence in
    the Kanechlor 300, 100 mg/kg diet) is not significant and the 2 higher
    dose levels of this product did not induce such changes. It can be
    concluded that Kanechlor 300 did not induce neoplasia at dietary
    levels of up to 1000 mg/kg over a 52-week period, in this study. At
    the 1000 mg/kg level, Kanechlor 300 did produce other evidence of
    chronic liver toxicity including oval cell and bile-duct
    proliferation, fatty changes, and cellular hypertrophy and, possibly,
    cholangiofibrosis (2/15). In the case of Kanechlor 500 (100 mg/kg
    diet), essentially the same picture was obtained. In this group, 3/25
    cases of nodular hyperplasia were found. In the case of Kanechlor 400
    with the dietary levels of 100 and 1000 mg/kg, 2/16 and 3/10 of the
    animals had nodular hyperplasia in the liver, respectively.

    In a preliminary study on Sherman rats (10 animals of each sex/dose),
    Aroclor 1254 at dietary levels of 0, 20, 100, or 500 mg/kg, and
    Aroclor 1260 at levels of 0, 20, 100, 500, or 1000 mg/kg, for 8
    months, did not give neoplastic nodules or hepatocellular carcinoma.
    At 500 mg/kg, Aroclor 1254 produced adenofibrosis in 10/10 male
    animals and at 100 and 500 mg/kg, in 7/10 and 9/10 females,
    respectively. This change was only seen in 2/10 male and 4/10 female
    animals with Aroclor 1260. The authors stated that hepatocellular
    adenofibrosis is a persistent progressive lesion that consists of a
    marked proliferation of fibrous tissue and epithelial glandular cells
    that are well differentiated in mice, but appear atypical in rats
    (Kimbrough et al., 1972).

    A group of 200 female Sherman rats was given Aroclor 1260 at an
    average dietary level of 100 mg/kg (range, 70-107 mg/kg), for 21
    months. The PCB intake declined from 11.6 mg/kg per day during the
    first week to 6.1 mg/kg at 3 months and to 4.3 mg/kg body weight per
    day, later on. The control group also comprised 200 rats. The survival
    rate and the food intake were not affected and no treatment-related
    signs of toxicity were observed. Body weight gain was decreased from 3
    months after the onset of the exposures. Hepatocellular carcinomas
    were present in the liver of 26/184 (14%) exposed rats and 1/173
    (0.58%) control rats. The livers of most of the remaining exposed rats
    (144/184) showed hyperplastic nodules, while none were found in
    control rats. A total of 182 exposed rats and 28 control rats had
    livers with foci or areas of cytoplasmic alteration. A few livers of
    exposed rats showed adenofibrosis. No induction of tumours in other
    organs and no metastases from the liver tumours were found (Kimbrough
    et al., 1975).

    Calandra (1976) reported the findings from several long-term studies,
    performed by a commercial laboratory for Monsanto. (These studies have
    never been published). In these studies, 1000 rats were divided into
    10 groups of 100 animals (50 of each sex) and 9 of the groups were
    exposed to Aroclors 1242, 1254, or 1260 at dietary levels of 1, 10, or
    100 mg/kg diet. Apparently, 5 animals of each sex were sacrificed at
    3, 6, and 12 months with about 35 animals killed at the end of the
    2-year studies. In the animals sacrificed early, only one nodular
    hyperplasia was observed and it was in the group fed 100 mg Aroclor
    1260/kg for 12 months. Mortality in these studies was high and
    approximately one-third of the 105 animals anticipated to be exposed
    for 2 years at each dietary level died. Hepatomas were observed in
    7/25 livers from animals fed 100 mg Aroclor 1260/kg, in 4/26 fed
    Aroclor 1254, 3/19 fed Aroclor 1242, and only in 1/168 animals
    receiving the 1-10 mg/kg diets. Nodular hyperplasia was twice as
    prevalent as hepatomas in the high-dose animals, particularly in the
    Aroclor 1254 group (Harbison et al., 1987).

    In a limited study, groups of 24 Fischer 344 rats of each sex received
    a diet containing Aroclor 1254 at 0, 25, 50, or 100 mg/kg diet
    (equivalent to 0, 1.2, 2.5, and 5 mg/kg body weight, respectively) for
    104-105 weeks. The survival rate decreased with a dose-related trend
    in male, but not in female rats (92, 83, 58, and 46%, respectively).
    From 10 weeks of exposure onwards, the body weight gain of all rats,
    except that of the low-dose males, decreased. In the groups receiving
    the 2 highest dose levels, alopecia, facial oedema, exophthalmos, and
    cyanosis were observed. Foci of hepatocellular alterations, which were
    dose-related, were found at all dose levels, but not in the controls.
    The incidence of non-neoplastic hyperplastic nodules in male rats with
    25, 50, or 100 mg/kg diet was 5/24, 8/24, and 12/24 and in female rats
    6/24, 9/22, and 17/24, respectively (US EPA, 1980). None was found in
    the controls. Hepatocellular adenoma and carcinomas were found in 1/24
    males and 1/24 females on the 50 mg/kg diet and in 3/24 males and in
    2/24 females on the 100 mg/kg diet. Non-neoplastic liver lesions
    included degenerative changes and aggregates of macrophages with
    crystalline cytoplasmic structures and pigment granules. An apparently
    dose-related increase in the incidence of intestinal metaplasia was
    observed in both sexes and 0, 1, 3, and 2 adenocarcinomas located in
    the pyloric region of the glandular stomach were found at 0, 25, 50,
    and 100 mg/kg diet, respectively. Morgan et al. (1981) and Ward (1985)
    reexamined the NCI (1978) data with respect to gastric
    adenocarcinomas, hepatocellular adenomas, and carcinomas. Morgan et
    al. (1981) found incidences of focal stomach lesions, mostly
    metaplasia, of 6, 10, 17, and 35% in rats receiving 0, 25, 50, or
    100 mg Aroclor 1254/kg diet, respectively. Adenomas were found in 6
    treated rats. When compared with the incidences of stomach
    adenocarcinomas in historical controls (1/3548), the incidence of
    6/144 was statistically significantly increased (NCI, 1978; Morgan et
    al., 1981; Ward, 1985).

    Groups of 70 Sprague-Dawley rats of each sex received a diet
    containing Aroclor 1260 in corn oil at a concentration of 100 mg/kg
    diet (equivalent to 5 mg/kg body weight) for 16 months and 50 mg/kg
    diet (equivalent to 2.5 mg/kg body weight) for an additional 8 months.
    All surviving rats received a basal diet from month 25 to month 29.
    The control group comprising 63 rats of each sex, received the basal
    diet with corn oil for 18 months and the basal diet alone for an
    additional 5 months. All surviving rats received the basal diet from
    the 25th month to the 29th month. Data on growth were not available.
    Groups of 2 control or 3 PCB-treated rats of each sex were partially
    hepatectomized at 1, 3, 6, 9, 12, 15, and 18 months. At 24 months, a
    similar group was sacrificed; after 29 months all remaining animals
    were sacrificed. The mortality rate was not affected by the exposure.
    In the livers of exposed rats, centrolobular hypertrophy was apparent
    at 1 month, foci at 3 months, and areas of cellular alteration after 6
    months, neoplastic nodules after 12 months, trabecular carcinomas
    after 15 months, and adenocarcinomas after 24 months. Metastases in
    the lung were not found. In exposed rats that survived 18 months or
    longer, hepatocellular carcinomas were present in 43/47 females and in
    2/46 males, but were absent in 81 controls. Simple and cystic
    cholangioma at 18 and 23 months, respectively, and adenofibrosis at 22
    months were present in the treated rats (Norback & Weltman, 1985).

    Rao & Banerji (1988b) fed 3 groups of 32 weanling Wistar rats a
    protein diet containing 0 (coconut oil), 50, or 100 mg Aroclor 1260/kg
    diet, for 120 days. The incidence of neoplastic nodules in the liver
    was 0/32, 24/32 and 16/32, respectively. Adenofibrosis was also
    observed in the treated animals.

    In order to exclude possible effects of dibenzofurans and to
    investigate the effect of the degree of biphenyl-chlorination, groups
    of 152 and 144 male Wistar rats were exposed to Clophen A30 and
    Clophen A60, respectively, not containing detectable quantities
    (detection limit not stated) of dibenzofurans, for 800 days at a dose
    of 100 mg/kg diet. A group of 139 rats received a control diet. After
    800 days, randomly selected rats were killed daily, until all
    survivors had been examined by day 832. The survival rate of the
    remaining exposed rats was increased by day 800. In rats autopsied by
    day 800 and in rats autopsied later, the incidence of hepatocellular
    carcinoma was increased by exposure to Clophen A60 (9/129 and 52/85,
    respectively) relative to controls (0/131 and 1/53, respectively), but
    not by exposure to Clophen A30 (1/138 and 3/87, respectively). There
    was a marked trend from foci of hepatocellular alteration to
    neoplastic nodule to carcinoma with increasing time and degree of
    biphenyl chlorination. Controls mainly showed foci. Non-neoplastic
    liver lesions with increased incidences of bile duct hyperplasia were
    found in rats receiving Clophen A30 and A60, and cysts in rats
    receiving Clophen A60. The incidence of adenofibrosis in the liver was
    decreased, compared with that in the controls, in all exposed rats
    that were killed after exposure (Schaeffer et al., 1984).

    8.5.2   Tumour promotion/anticarcinogenic effects

    (a) Mice

    Tatematsu et al. (1979) studied the effects of inducers of liver
    microsomal enzymes on the induction of hyperplastic liver nodules by
     N-2-fluorenylacetamide (2-FAA) in male F344 rats. The rats were fed
    a diet containing 200 mg 2-FAA/kg diet for 2 weeks and then given 500
    or 1000 mg Kanechlor 500/kg diet for the following 8 weeks. Partial
    hepatectomies were performed at the end of the third week of the
    study. Kanechlor 500 in the dose levels applied showed a promoting

    Ito et al. (1973) and Nagasaki et al. (1974, 1975) exposed groups of
    20-38 male dd mice to diets containing Kanechlor 400 or 500 at 100 or
    250 mg/kg for 24 weeks. Control groups consisted of 20 mice. Combined
    exposure to Kanechlor 500 at 100 and 250 mg and either 50, 100, or
    250 mg alpha- or -BHC (hexachlorocyclohexane)/kg diet enhanced the
    development of nodular hyperplasia and hepatocellular carcinomas. A
    combination of Kanechlor 500 and gamma-hexachlorocyclohexane did not
    produce tumours. Dosing with Kanechlor 500 alone at a dietary level of
    100 and 250 mg/kg, and - or gamma-hexachlorocyclohexane at dietary
    levels of up to 500 mg/kg did not produce tumours. However,
    alpha-hexachlorocyclohexane (250 mg/kg diet) produced 10/38
    hepatocellular carcinomas and 30/38 hyperplastic nodules.

    In another study, both inhibition and promotion were observed
    following transplacental and transmammary exposure of mice to PCBs
    prior to, or simultaneously with, exposure to dimethylnitrosamine
    (DMNA). In this study, Aroclor 1254 was administered intraperitoneally
    to female Swiss CD-1 mice on the 19th day of gestation, at a dose of
    500 mg/kg body weight. Groups of 17-31 sucklings of these mice and of
    controls were then treated intraperitoneally with DMNA on postnatal
    day 4 or 14, or remained untreated. The progeny were killed at 28
    weeks or at 18 months of age. Aroclor 1254 exposure decreased the
    incidence of tumours in the liver and lung, induced by DMNA
    administered at postnatal day 14. The average numbers of lung tumours
    per mouse were also decreased. However, Aroclor 1254 increased the
    liver tumour-bearing mice with extensive DMNA-initiated liver tumours
    at 18 months of age, especially when DMNA was administered on
    postnatal day 4. Mice exposed only to Aroclor 1254 did not show tumour
    incidences higher than those of controls (Anderson et al., 1983). The
    authors also reported that 2 higher chlorinated biphenyls,
    2,4,5,2',4',5'-hexachlorobiphenyl and 2,3,5,2',3',5'-hexa-
    chlorobiphenyl, were the dominant congeners persisting in the tissues
    of the PCB-treated mice. It should be noted that only the more highly
    chlorinated PCBs (>50% by weight) were reported as promoters of
    hepatocarcinogenesis in rodents. The promoting activities of the lower
    chlorinated PCBs have not been determined so far.

    (b) Rat

    As shown in Table 49, administration of PCBs to rats after exposure to
    several initiating agents promoted the development of neoplastic
    lesions of the liver. In these studies, treatment of the rats with a
    control diet of PCBs alone did not usually lead to neoplastic changes
    in the liver. Preston et al. (1981) compared the promoting effect of
    exposure of rats to Aroclor 1254 with that of exposure to Aroclor 1254
    from which the polychlorinated dibenzofuran moieties had been removed,
    and did not find any significant differences. The results of this
    study suggested that the promoting effect of commercial PCBs cannot be
    ascribed entirely to the presence of chlorinated dibenzofurans.
    Tatematsu et al. (1979) showed a dose-effect relationship in the
    observed promoting effect of Kanechlor 500.

    Inhibition, rather than promotion, of liver neoplasms was observed
    when female Donryu rats were exposed to Kanechlor 400 preceding, or
    simultaneously with, exposure to 3'-methyl-4-dimethylaminoazobenzene
    (3'-Me-DAB) (Kimura et al., 1976). Simultaneous dietary exposure of
    groups of 16-24 male Sprague-Dawley rats to Kanechlor 500, at a
    dose-level of 500 mg/kg (25 mg/kg body weight per day), and to
    3'-Me-DAB, or  N-2-fluorenylacetamide (2-FAA), or diethylnitrosamine
    (DENA), or combinations of these carcinogens, for 20 weeks, showed
    almost complete inhibition of neoplastic nodules and hepatocarcinomas
    (Makiura et al., 1974). It should be noted that the results of this
    study may be influenced by the extreme toxicity, as shown by the
    weight records, especially when the substances were combined.

    The antitumour activity of Aroclor 1254 was demonstrated in male
    Sprague-Dawley rats, inoculated with Walker 256 tumour cells. Groups
    of 16 rats received PCB doses of 50, 100, or 200 mg/kg body weight for
    2 weeks, once in 2 days, starting on the day of tumour cell injection.
    A group of 16 rats did not receive PCBs. The inhibition of tumour
    growth and transplantability were dose-related (Kerkvliet & Kimeldorf,
    1977). The antitumour activity of Phenoclor DP5 was observed in groups
    of 20 female Swiss mice inoculated with Ehrlich's tumoral ascites
    liquid, after receiving PCBs in the diet at levels of 0, 10, 50, or
    250 mg/kg, for 120 days (Keck, 1982).

    Nishizumi (1980) showed that placentally transferred PCBs inhibited
    diethylnitrosamine (DENA)-induced liver tumours. Groups of ten,
    10-week-old female Wistar rats were treated with 40 or 200 mg
    Kanechlor 500/kg body weight, by gavage, on days 5, 10, and 15 of
    gestation. One F1 offspring from each litter was killed for
    quantification of liver PCBs. The remaining F1 offspring were exposed
    to 50 mg DENA/litre drinking-water, continuously for 5 weeks. At 16,
    20, and 24 weeks after the beginning of the DENA exposure, 6-8 rats of

        Table 49.  Promotion of liver neoplasms in rats exposed to PCBs after treatment with carcinogenic substances

    Strain    Sex      Group size    Carcinogenic   PCBs             Doseb         Exposure    Neoplasms      Reference
                       substancesa                                   (mg/kg diet)  period      promoted by
                                                                                   (weeks)     PCBs

    Donryu    female   25            3'-Me-DAB      Kanechlor 400    400           26          carcinoma      Kimura et al.

    Wistar    male     20-24         DENA           Kanechlor 500    2  15        4           nodules        Nishizumi (1976)
                                                                     mg/week                   carcinoma

    F 344     male     15-16         2-FAA          Kanechlor 500    500 and       8           hyperplasia    Tatematsu et al.
                                                                     1000                      nodules        (1979)

    F 344     male     20            EHEN           Kanechlor        500           32          carcinoma      Hirose et al.

    Sprague-  male     40            DENA           Aroclor 1254c    100           18          carcinoma      Preston et al.
    Dawley                                                                                                    (1981)

    a  3'-Me-DAB = 3'-methyl-4-dimethylaminoazobenzene; DENA = diethylnitrosamine; 2-FAA = N-2-fluorenylacetamide;
       EHEN = N-ethyl-N-hydroxyethylnitrosamine.
    b  Unless otherwise specified.
    c  With and without PCDFs.

    each sex from each treatment group were killed and examined
    histologically. A significant reduction in tumour incidence occurred
    only in male offspring in the 200-mg group. The liver-PCB values in
    28-day-old F1 mice were < 1, 18  7, and 360  30 mg/kg tissue for
    the controls, 40, and 200 mg/kg groups, respectively. It was suggested
    that placental transfer of PCBs protected the treated rats from
    DENA-induced liver tumours.

    The inhibitory effect of PCBs on tumour initiators following
    simultaneous exposure has been explained by an enhanced metabolism of
    the initiator by PCB-induced mixed function oxygenase.

    Tests for putatively preneoplastic enzyme-altered foci in the liver of
    rats have shown a dose-related increase in the promotion of such foci
    by intraperitoneal exposure to Aroclor 1254 in tricaprylin or by oral
    exposure to Clophen A50 in olive oil, following oral administration of
    diethylnitrosamine (Deml & Oesterle, 1982; Pereira et al., 1982;
    Oesterle & Deml, 1983).

    The highest oral dose of Clophen A50 not enhancing the number and area
    of enzyme-altered loci in female Sprague-Dawley rats, treated for 11
    weeks with doses of 0, 0.1, 0.5, 1, 5, or 10 mg/kg body weight after
    initiation by a single dose of 8 mg/kg body weight of
    diethylnitrosamine, was 0.5 mg/kg body weight (Deml & Oesterle, 1987).

    8.5.3  Initiation, promotion, and other special studies on individual

    2,5,2',5'-Tetrachlorobiphenyl and its metabolite 2,5,2',5'-
    tetrachlorobiphenyl-3,4-oxide were tested in a pulmonary tumour
    induction assay with intraperitoneally injected A/T mice of both
    sexes, and in a two-stage skin carcinogenicity assay with
    dermally-exposed female SENCAR mice. Pulmonary adenomas and skin
    papillomas were not induced (Preston et al., 1985).

    Female Harlan-Sprague-Dawley rats received 2,4,2',4'-
    tetrachlorobiphenyl or 2,5,2',5'-tetrachlorobiphenyl at 100 mg/kg
    diet for 28 weeks, 1 week after oral exposure to diethylnitrosamine
    (DENA). Both congeners showed a promoting effect on the development of
    foci of hepatocellular alteration. The effect was approximately
    10-fold greater in rats receiving 2,4,2',4'-tetrachlorobiphenyl
    (Preston et al., 1985).

    The effects of PCB mixtures and selected congeners have also been
    investigated by Hayes et al. (1985, 1986) using the resistant
    hepatocyte model developed by Farber and coworkers (Solt & Farber,
    1976; Tsuda et al., 1980; Farber, 1984a,b, 1986). The ability of
    2,4,2',4' and 2,5,2',5'-tetrachlorobiphenyl, 2,4,5,2',4',5'-
    hexachlorobiphenyl, and a mixture of PCB-congeners (the composition of

    which resembled that ascertained in human breast milk) to initiate
    enzyme-altered hepatocellular nodules was investigated in
    proliferating hepatocytes of neonatal or partially hepatectomized
    adult rats (the PCB-congeners did not contain detectable levels of
    dibenzofurans or dioxins). Neonatal rats were exposed 3 times in 3
    weeks and adult rats once. After several weeks, the rats received a
    selection regimen of 2-acetylaminofluorene followed by partial
    hepatectomy (neonates) or necrotizing carbon tetrachloride (adults).
    None of the PCB exposures generated nodules in contrast to known
    initiators (Hayes et al., 1985).

    Subsequent studies by Hayes et al. (1986) using the afore-mentioned
    compounds and 3,4,3',4'-tetrachlorobiphenyl (a typical MC-type inducer
    of cytochrome P-450-dependent monooxygenases) showed that these PCBs
    (50 mol/kg body weight), given 10 days after a dose of the initiator,
    DENA, and 7 days before 2-AAF, all reduced the size of the
    2-AAF-selected gamma-glutamyltranspeptidase-positive nodules. These
    results show that, in contrast to the previous studies, PCBs also
    exhibit "anti-promoting" activities in this Farber-model, which
    utilizes 2-AAF as a mito-inhibitory toxin.

    8.5.4   Skin carcinogenicity

    Aroclor 1254 (100 g/mouse) administered 18 h prior to the initiator,
    7,12-dimethylbenz (a)anthracene (DMBA), significantly decreased the
    incidence of papilloma formation in female Charles-River CD-1 mice.
    2,4,5,2',4',5'-Hexachlorobiphenyl (625 g/mouse), a PCB congener that
    resembles phenobarbital in its mode of induction of drug-metabolizing
    enzymes, did not act as an anticarcinogen, whereas 3,4,3',4'-
    tetrachlorobiphenyl was more active than Aroclor 1254 as an
    inhibitor. 3,4,3',4'-Tetrachlorobiphenyl resembled methylcholanthrene
    in their mode of cytochrome P-450 induction (Parkinson et al., 1983)
    and decreased the number of DMBA-initiated papillomas/mouse. Although
    the PCB treatment did not modulate the incidence of papillomas caused
    by benzo (a)pyrene, it was suggested that the anticarcinogenic
    effects of PCBs on mouse skin tumours initiated by DMBA were due to
    altered metabolism and DNA binding of the carcinogen by the
    PCB-induced skin monooxygenases (DiGiovanni et al., 1979).

    Berry et al. (1978, 1979) studied the tumour-promoting activity of
    Aroclor 1254 in groups of 30 female CD-1 mice initiated with
    0.2 mmol/litre (equivalent to 51 g) DMBA. One week later, a positive
    control group received 2 g tetradecanoylphorbolacetate (TPA) and an
    experimental group received 100 g Aroclor 1254 in acetone. The TPA
    and Aroclor applications were made twice weekly for 30 weeks. The TPA
    promotion resulted in 92% of the animals developing papillomas, while
    none developed in the Aroclor-treated animals. It was concluded that
    Aroclor 1254 was not a skin tumour promoter at the dose used in this
    study. Inhibition of skin papilloma was observed when Aroclor 1254 was
    administered 18 to 72 h prior to DMBA treatment and promotion by TPA.

    Poland et al. (1982) used HRS/J hairless mice to study the tumour
    promotion activity of Aroclor 1254 in combination with  N-methyl-
     N'-nitro- N-nitrosoguanidine (MNNG). Twenty female mice per group
    received a single administration of MNNG (5 micromol in acetone) or
    acetone alone applied on the skin, and were then treated topically,
    twice weekly, with 1 mg Aroclor/mouse, dissolved in acetone, for 20
    weeks. No tumours were induced in the control group; in the
    MNNG-treated mice, 4/19 had papillomas. It was concluded that Aroclor
    had a weak promoting effect.

    8.5.5   Appraisal

    In summarizing the potential carcinogenic activity of PCBs, it is
    perhaps more informative to express it in terms of what is known about
    the mechanisms of chemical carcinogenicity (Hayes, 1987). In other
    words, what the evidence is to support the carcinogenicity of PCBs
    through genotoxic/initiating, cocarcinogenic, promotional/
    antipromotional, and progressional activities. There is no evidence to
    support genotoxic activity for PCBs and, in  in vitro studies,
    evidence for initiating activity through direct interactions with DNA
    is weak. Poor initiating activity is consistent with the demonstrated
    lack of mutagenic activity in various short-term tests. There is
    evidence to suggest that PCBs can potentiate the activity of known
    carcinogens (or act as cocarcinogens) in  in vitro systems, but the
    opposite result is often seen in  in vivo studies, suggesting that
    certain protecting enzyme systems present in the intact systems are
    absent in the  in vitro systems. There is a substantial body of
    evidence to support the promotional activity of PCBs, particularly the
    more highly chlorinated ones, in rodent liver, and this activity may
    depend on the sequence in which the chemicals are administered in
    experimental animal studies. In addition, promotional activity is
    correlated closely, but not consistently, with the induction of MFO
    activity. The hyperplastic effect (stimulation of cell proliferation)
    of PCB inducers could promote preneoplastic growth. This type of
    activity may possibly involve a threshold, suggesting that it may not
    be a factor in low-level exposures to PCBs. The anticarcinogenic
    activity of PCBs may also depend on the sequence of events in
    experimental animal studies, and it may be related more to the
    antipromotional properties of PCBs, possibly functioning as
    protectants of mito-inhibitory toxicity. The interpretation of the
    available animal data involving the commercial PCB mixtures is often
    complicated by lack of information concerning the presence and
    contribution of chlorinated dibenzofuran impurities, as well as
    variations in congener composition to toxicity. In structure-activity
    terms, a key factor in determining promotional activity appears to be
    the degree of chlorination, which may reflect increased resistance to

    metabolism and elimination and possibly also higher degrees of
     ortho-substitution among the congeners present.  Ortho-substituted
    PCBs (possibly acting as persistent PB type inducers) have been shown
    to be effective tumour promoters, and, at least in the case of the
    closely related PBBs, it has been shown that a non-toxic and
    non-promoting dose of a non- ortho-substituted congener in
    combination with a promoting dose of a highly  ortho-substituted
    congener has a synergistic effect. This result may explain why the
    mixtures can have greater promoting ability than the individual
    congeners involved. This result might also suggest multiple pathways
    for promoting activity, possibly involving the Ah receptor as well as
    the putative receptor for phenobarbital. The possibility that PCBs
    might promote carcinogenesis in tissues, other than liver, in animals
    exposed to various tissue-specific, initiating agents, needs to be
    addressed. Nevertheless, the potential for human liver cancer from
    exposure to PCBs cannot be reliably predicted from animal studies.
    Overall, there is reason to exercise caution in extrapolating the
    available animal data on the carcinogenic potential of PCBs for

    8.6  Special studies: target-organ effects

    The lesions commonly introduced in animals after acute, short-term, or
    long-term administration/application of PCB mixtures and/or individual
    congeners concern the liver, skin, immune system, reproductive system,
    oedema at various sites, as well as disturbances of the
    gastrointestinal tract and the thyroid gland.

    8.6.1  Liver  PCB mixtures

    The toxic effects of PCBs result both directly and indirectly from
    their presence in certain organs, such as the liver, where they
    induce, in various degrees, a variety of liver enzymes. Some of these
    enzymes are active in the metabolism of the PCBs themselves, while
    others involve activation, deactivation, detoxication, etc., of other

    In itself, induction of enzymes by xenobiotics does not represent a
    toxic manifestation, rather it is the ordinary response to such
    foreign chemicals, which results generally in their detoxication and
    ultimate modification, enabling them to be excreted from the organism.
    In this sense, the response of the liver to such compounds is a
    biological protective mechanism. Since some compounds, including PCBs,

    are capable of inducing not only enzymes that result in their own
    detoxication, but others as well, and the level of enzyme induction
    may be high enough to cause liver pathology, a line cannot be drawn
    that clearly separates a normal biological function from a toxic
    manifestation. Superimposed on this situation is the direct toxic
    action of the compounds on liver tissue, because of the properties of
    the parent compound or its metabolites.

    An enlarged liver and increased absolute and relative liver weights
    are commonly reported as gross effects of PCB administration. The
    lowest-observed-effect levels, in studies on different rat strains,
    exposed to a diet containing Aroclor 1254, for a (dose-related)
    increase in relative liver weights, vary between 20 and 100 mg/kg diet
    (equivalent to 1 and 5 mg/kg body weight, respectively) (Kimbrough et
    al., 1972; Bruckner et al., 1974; Burse et al., 1974; Grant et al.,
    1974; Allen et al., 1976; Zinkl, 1977; Hinton et al., 1978; Kasza et
    al., 1978b; Jonsson et al., 1981; Baumann et al., 1983).

    The liver hypertrophy is microscopically visible as enlarged,
    pleiomorphic hepatocytes, sometimes multinucleated or with enlarged
    nuclei. A liver alteration observed by many investigators after
    exposure of rats to various PCB-mixtures is fatty degeneration,
    characterized by fat vacuolation and/or a foamy appearance of the
    cytoplasm. Ultrastructurally, an increase in the number and size of
    cytoplasmic lipid droplets and liposomes (membrane-associated lipid
    droplets) can be observed. Fatty degeneration of the liver was already
    observed after exposure of male Sprague-Dawley rats to a diet
    containing 5 mg Aroclor 1242/kg diet (equivalent to 0.25 mg/kg body
    weight) for 2-6 months (Bruckner et al., 1974) and after exposure of
    male Holtzman rats to a diet containing 5 mg Aroclor 1254/kg for 5
    weeks (Kasza et al., 1978b). Chu et al. (1977) found a comparable
    effect on the liver with >20 mg Aroclor 1254 or 1260/kg diet for 28
    days. Another characteristic change is the appearance of eosinophilic,
    lamellar, cytoplasmic inclusions (Kimbrough et al., 1972; Kasza et
    al., 1978b) or "hyaline-like material" (Grant et al., 1974). Electron
    microscopy revealed that these changes corresponded to concentric
    laminated membranes of smooth endoplasmic reticulum ("whorls"; "myelin
    figures") (Vos & Beems, 1971; Kimbrough et al., 1972; Allen et al.,
    1976; Kasza et al., 1978b; Jonsson et al., 1981). Proliferation of the
    smooth endoplasmic reticulum and a decrease in rough endoplasmic
    reticulum are commonly observed in rats exposed to dietary levels of
    PCB mixtures that also induce fatty degeneration. Kasza et al. (1978b)
    further observed a marked proliferation of Golgi condensing vesicles
    containing lipoprotein in the livers of male Holtzman rats fed Aroclor
    1254, for 5 weeks, at 5 mg/kg diet. A decreased number of these
    vesicles was seen at 50 and 500 mg/kg diet (equivalent to 2.5 and
    25 mg/kg body weight, respectively), together with a marked increase

    in the smooth endoplasmic reticulum and lysosomes. The above decrease
    in Golgi vesicles was also observed in Sprague-Dawley rats by Hinton
    et al. (1978). The proliferative changes of the endoplasmic reticulum
    are closely related to the observed induction of microsomal enzymes,
    as discussed in section Atypical mitochondria (Burse et al.,
    1974; Kasza et al., 1978a,b), and single cell and focal necrosis
    (Grant et al., 1974; Allen et al., 1976; Jonsson et al., 1981; Baumann
    et al., 1983) have also been described.

    Hypobilirubinaemia was produced in rats by Bastomsky et al. (1975),
    who investigated the mechanism by administering daily intraperitoneal
    injections of Aroclor 1254 (25 mg/kg body weight, in corn oil) to
    female rats for 4 days, before measuring bilirubin glucuronide
    formation by hepatic microsomes  in vitro. PCB treatment was not
    effective in increasing UDP-glucuronosyltransferase (EC
    activity. Serum bilirubin levels in Gunn rats were also significantly
    decreased by PCB treatment; the rats are genetically deficient in
    UDP-glucuronosyltransferase ( activity.

    The fluorescing of livers on exposure to UV radiation, consistent with
    the presence of porphyrin, and accumulation of brown pigment, positive
    for iron, especially in Kupffer cells and perivascular macrophages
    have also been reported (Kimbrough et al., 1972; Burse et al., 1974;
    Zinkl, 1977; Jonsson et al., 1981). The changes described in the
    livers of mice (Nishizumi, 1970) and rabbits (Koller & Zinkl, 1973),
    following exposure to PCB mixtures, are comparable with those in rats.  Individual congeners

    The effects of chlorination and the chemical composition of PCBs, with
    regard to the dose-effects relations in liver toxicity after
    short-term exposure, are indicated by the data of Biocca et al.
    (1981). In this study, hepatotoxic effects were observed in mice after
    5 weeks of maintenance on diets containing 0.3 mg of 3,4,5,3',4',5'-
    hexachlorobiphenyl, while similar effects were observed only after
    30 mg of 2,4,5,2',4',5'-hexachlorobiphenyl and 100 mg of 2,4,6,2',4',6'-
    hexachlorobiphenyl/kg diet. No effects were found with 300 mg of
    2,3,6,2',3',6'-hexachlorobiphenyl/kg diet. Similar dependence of liver
    toxicity on the chemical composition of the PCB mixture would be
    anticipated following long-term exposure in mice and other species.

    8.6.2  Enzyme induction  Effects on liver enzymes of PCBs

    Proliferation of the smooth endoplasmic reticulum is a common
    observation in the liver cells of experimental animals following
    exposure to PCB mixtures. This effect is accompanied by an increase in
    microsomal protein and the induction of cytochrome P-450, cytochrome
    P-448, and drug-metabolizing enzymes, including the microsomal

    monooxygenases (EC, epoxide hydrolases (EC,
    UDP-glucuronosyltransferases (EC, NADPH-cytochrome c
    reductase (EC and esterases (EC, and the cytosolic
    glutathione  S-transferase (EC The subject has been
    reviewed by Safe (1984).

    The spectral, enzymatic, and electrophoretic properties of the
    microsomal enzymes, induced by Aroclor 1248, 1254, 1260, and Kanechlor
    400, are consistent with the inducing properties of both the
    phenobarbital (PB) and 3-methylcholanthrene (MC) classes of inducers.
    They induce both cytochromes P-450 and P-448 and associated enzymes
    (Alvares & Kappas, 1977; Goldstein et al., 1977; Yoshimura et al.,
    1978; Iverson et al., 1982; Lashneva & Tutelyan, 1984; Khan et al.,
    1985; Tutelyan et al., 1986). Aroclor 1016, administered to
    Sprague-Dawley rats at 50 mg/kg per day for 4 days, intraperitoneally,
    elicited a barbiturate type of inducing effect on the hepatic
    microsomal oxidative enzyme system. Aroclor 1016 caused increases in
    liver cytochrome P-450 content, microsomal protein, and its
    ethylmorphine  N-demethylase activity. It did not induce cytochrome
    P-448 in liver microsomes.

    A dose-related induction of hepatic and, in some cases, extrahepatic
    microsomal enzymes was observed in several animal species including
    the rat, rabbit, mouse, ferret, guinea-pig, hamster (Safe, 1984), mink
    (Shull et al., 1982; Aulerich et al., 1985) and monkey (Iverson et
    al., 1982). Distinct interspecies variations have been demonstrated.
    For example, while 6 daily intraperitoneal doses of 25 mg of Aroclor
    1254/kg body weight caused a potent induction of benzo (a)pyrene
    hydroxylase in adult, male Sprague-Dawley rats, no, or a minimal,
    induction of this monooxygenase was observed in adult, male Swiss mice
    after 4 daily intraperitoneal doses of 50 mg/kg body weight and in
    male New Zealand rabbits after 2 intraperitoneal doses of 100 mg/kg
    body weight on days 1 and 4 (Alvares et al., 1982). Furthermore, when
    comparing the inducing potency of Aroclor 1242 in the mink and the
    genetically related ferret, Shull et al. (1982) measured a greater
    induction of cytochrome P-448 and MC-type monooxygenases and no toxic
    effects in the ferret at a dosing regime that resulted in toxic
    effects in the mink (100 mg/kg body weight on day 1,200 mg/kg body
    weight on day 5, sacrifice on day 10). The authors considered the
    observed induction moderate in both species compared with that
    observed in the rat. Earlier, it had been found that male
    Sprague-Dawley rats were indeed more sensitive than ferrets with
    respect to the inducing effect of Aroclor 1254 following a single
    intraperitoneal dose of 500 mg/kg body weight, though the responses of
    both species were comparable qualitatively (Lake et al., 1979).

    Moreover, pretreatment of rats with Aroclor 1254 resulted in the
    induction of microsomal cytochromes P-450 c, P-450 d (MC-inducible),
    P-450 b and P-450 e (PB-inducible) (Ryan et al., 1979a,b, 1982). In
    general, the extent of induction of microsomal enzymes by PCB-mixtures
    increased with increasing chlorine content up to 54%. The effect has
    also been demonstrated with single pure PCBs administered orally
    (Ecobichon & Comeau, 1975). The results are summarized in Table 50 and
    show a greater degree of enzyme induction with the higher chlorinated
    compounds (see section

        Table 50.  Stimulation of microsomal enzyme activity by single chlorinated biphenylsa

                           Hepatic microsomal enzyme activity
    Chlorine               O-Demethylation   N-Demethylation   Aniline          Nitro-
    substituents                                               hydroxylation    reduction

    4                          0                 0                 0               0
    2,2'                       0                 +                 +               0
    2,4'                       0                 0                 +               0
    4,4'                     + +               + +               + +               +
    2,5,2',5'                  0                 +               + +               +
    2,4,2',4'                + +               + +               + +               0
    2,4,5,2',4',5'           + +               + +               + +             + +
    2,3,5,2',3',5'           + +               + +               + +             + +
    2,4,6,2',4',6'           + +               + +               + +             + +
    2,3,4,5,2',3',4',5'      + +               + +               + +             + +

    a  From: Johnstone et al. (1974).
       0 = No activity.
       + = Slight activity.
       + + = Marked activity.

    Litterst et al. (1972) exposed groups of 6 male Osborn-Mendel rats to
    Aroclors 1242, 1248, 1254, or 1260 in the diet, at concentrations of
    0, 0.5, 5.0, 50, or 500 mg/kg diet, for 4 weeks. Increased microsomal
    nitroreductase and demethylase activities occurred at 0.5 mg/kg or
    more, increased pentobarbital hydroxylation and increased relative
    liver weight occurred at 5.0 mg/kg or more, and increased liver
    triglycerides occurred at 50 mg/kg diet. An inducing activity similar
    to, or lower than, that of Aroclor 1254 has often been found for more
    chlorinated mixtures (Villeneuve et al., 1971a, 1972; Bickers et al.,

    1972; Chen & Dubois, 1973; Ecobichon & Comeau, 1974; Schmoldt et al.,
    1974; Sawyer et al., 1984). Aroclor 1016 and 1242, both containing 42%
    chlorine but differing in congener composition, showed qualitative and
    quantitative differences in inducing effects. For example, Aroclor
    1254 enhanced ethylmorphine  N-demethylase activity 3-fold, while the
    maximum increase produced by Aroclor 1016 was only 40%. The 2 Aroclors
    also differed in their induction of the various forms of cytochrome
    P-450 (Alvares et al., 1982). Adult, male Sprague-Dawley rats were
    administered, intraperitoneally, a dosage of 0 or 50 mg Aroclor 1016
    in corn oil/kg body weight, for 4 days. Aroclor 1016 was a potent
    inducer of  N-methylase but a poor inducer of benzo (a)pyrene
    hydroxylase. Administration of 100 mg Aroclor 1016 in corn oil/kg body
    weight per day to adult male New Zealand White rabbits, for 4 days,
    resulted in an increase in liver cytochrome P-450 activity and
    decreases in benzphetamine- N-demethylase and benzo (a)pyrene
    hydroxylase activities compared with the controls. 7-Ethoxycoumarine-
     O-deethylase and 7-ethoxyresorufin- O-deethylase activities were
    comparable with those in the controls (Ueng & Alvares, 1985).
    Therefore, it can be concluded that the degree and type of induction
    not only depends on the chlorine content of the mixture, but is also a
    function of the congener composition, as will be discussed further in
    the next section.

    Not only species specificity, but also marked tissue specificity has
    been observed. While Aroclor 1254 was found to be a potent inducer of
    cytochrome P-450 content and benzo (a)pyrene hydroxylase activity in
    the rat lung and liver (Alvares & Kappas, 1977), it caused a 46%
    decrease in cytochrome P-450 content, a 31% decrease in
    benzo (a)pyrene hydroxylase activity, and a 61% decrease in
    ethylmorphine  N-demethylase activity in the rabbit lung. Aroclor
    1254 caused induction of cytochrome P-450 and both enzymes in the
    kidneys of these rabbits, but benzo (a)pyrene hydroxylase was not
    induced in the liver (Alvares et al., 1982).

    The inducing effect of PCB mixtures on the monooxygenase system has
    been observed in the livers of both male and female rats (Chen &
    Dubois, 1973; Grant & Phillips, 1974), minks and ferrets (Lake et al.,
    1979; Shull et al., 1982), in the livers of pregnant rats (Alvares,
    1977) and rabbits (Villeneuve et al., 1971a), in the placenta of rats
    (Alvares & Kappas, 1975), in fetal and neonatal rat livers (Alvares &
    Kappas, 1975; Baker et al., 1977; Inoue et al., 1981; Jannetti &
    Anderson, 1981; Lashneva et al., 1987), in immature rat livers (Chen &
    Du Bois, 1973; Narbonne, 1980), and in mature and senescent rat livers
    (Birnbaum & Baird, 1978).

    The lowest-observed-adverse-effect levels and the no-observed-effect
    levels for enzyme induction, found in short-term diet studies on rats,
    are presented in Table 51.

    Bruckner et al. (1977) showed that when Aroclor 1254 was administered
    in the diet, at a level equivalent to 0.25 mg/kg body weight,
    microsomal enzyme activity was induced after 1 day of exposure.
    Narbonne (1980) found a significant induction with 0.1 mg Phenoclor
    DP6/kg body weight, given in the diet, after 3-5 days.

    After a few weeks of exposure to Aroclor 1254 or 1260, at low dietary
    levels of between 0.25 and 1.25 mg/kg body weight, a plateau in
    microsomal enzyme activity was reached that was maintained over
    several months of exposure (Chen & Du Bois, 1973; Grant et al., 1974;
    Bruckner et al., 1977). Following short-term exposure, dietary levels
    of between 5 and 25 mg/kg may stimulate microsomal enzymes for up to 4
    months (Grant et al., 1974; Bruckner et al., 1977). Levels of 0.05 or
    0.1 mg/kg body weight failed to produce any effects or the periods of
    induction at these levels were long (Grant et al., 1974; Chen & Du
    Bois, 1973).

    A marked induction of the liver monooxygenase system was observed in
    the offspring of female rats given a single oral dose of a mixture of
    PCBs (Sovol) (500 mg/kg body weight) on the 14th day of pregnancy. In
    one-day-old rats, an increase in cytochrome P-450 content associated
    with an increase in the cytochrome b 5 level, an increase in NADPH-
    cytochrome c reductase activity, an increased rate of aminopyrine-
     N-demethylation activity in microsomes and also increased
    3,4-benzo (a)pyrene hydroxylation, 7-ethoxycoumarin  O-deethylation,
    and NADPH-dependent lipid peroxidation were found. The activity of the
    cytochrome P-450 system in the young rats remained elevated during the
    early postnatal period (Lashneva et al., 1987).  Effects on liver enzymes of "biologically filtered" PCB

    Young Sprague-Dawley rats were administered a total of 38 oral doses
    of a PCB mixture in olive oil at 0, 0.25, 1.0, 4.0, 16.0, 64.0, 256,
    or 1025 g/kg body weight, twice a day, over 1 month; the mixture
    contained 55% chlorine and had a gas-chromatographic profile very
    similar to that of the congeners found in the breast milk of Japanese
    women. A dose-related induction of liver aminopyrine demethylase and
    benzo (a)pyrene hydroxylase (EC was found with doses of
    1.0 g/kg body weight or more. PCB-binding to liver microsomes was
    increased at doses of 4.0 g/kg body weight or more (Shimada & Ugawa,
    1978). In another study, 1-month-old, male Wistar rats received
    (intraperitoneally) doses of 0, 1, 10, 25, 50, or 100 mg/kg body
    weight of a reconstituted PCB-mixture in corn oil, containing average
    levels of 13 of the major congeners found in the breast milk of
    Japanese women (purity > 98.5%). Each dose was administered in 2
    portions on days 1 and 3. The same dose regimen of Kanechlor 500 was
    also tested. The rats were killed on day 6. The reconstituted PCB
    mixture and Kanechlor 500 caused dose-related increases in liver

        Table 51.  Microsomal enzyme induction by PCB-mixtures in rats

    PCB-mixture    Rat strain        Sex       Exposure      No-effect-level    Lowest-observed-         Reference
                                    male/      period        mg/kg body         adverse effect level,
                                    female                   weight             mg/kg body weight

    Aroclor 1016   Sprague-Dawley    male      3 weeksa        0.1              1                        Iverson et al. (1975)

    Aroclor 1242   Sprague-Dawley    male      3 weeksa        0.1              1                        Iverson et al. (1975)
                   Osborne-Mendel    male      4 weeks       < 0.025            0.025                    Litterst et al. (1972)
                   Sprague-Dawley    male      2-6 months    < 0.25             0.25                     Bruckner et al.
                                                                                                         (1974, 1977)

    Aroclor 1248   Osborne-Mendel    male      4 weeks       < 0.025            0.025                    Litterst et al. (1972)

    Aroclor 1254   Osborne-Mendel    male      4 weeks       < 0.025            0.025                    Litterst et al. (1972)
                   Wistar            male      2-8 months    < 0.1              0.1c                     Grant et al. (1974)
                   Wistar            male      2 weeks         0.25             0.5                      Den Tonkelaar &
                   Wistar            male      12 weeks        0.05             0.5                      van Esch (1974)
                   Sprague-Dawley    male      0-20 weeks      0.05             0.25                     Turner & Green (1974)
                   Holtzman          male      3 weeks       < 0.25             0.25                     Bruckner et al. (1977)
                                                                                                         Garthoff et al. (1977)

    Table 51.  (cont'd).

    PCB-mixture    Rat strain        Sex       Exposure      No-effect-level    Lowest-observed-         Reference
                                    male/      period        mg/kg body         adverse effect level,
                                    female                   weight             mg/kg body weight

    Aroclor 1260   Osborne-Mendel    male      4 weeks       < 0.025            0.025                    Litterst et al. (1972)
                   Holtzman          male      1-13 weeks    < 0.05             0.05                     Chen & DuBois (1973)
                                     female    1-13 weeks      0.05             0.25

    Phenoclor DP6  Sprague-          male      3 days        < 0.1              0.1                      Narbonne (1979, 1980)

    a  Daily dosing by gavage; the other studies are all diet studies.
    b  Immature rats (60-65 g).
    c  Significant effect after months 4 and 6, but not after months 2 and 8.

    benzo (a)pyrene hydroxylase activity that were 3.5 and 2.2 times the
    control value, respectively, at 1 mg/kg body weight. The ED50 of the
    reconstituted breast milk PCB mixture for the induction of rat hepatic
    microsomal aryl hydroxylase (AHH) was 7 times lower than that of
    Kanechlor 500. The authors concluded that the increased potency of the
    breast milk-PCB mixture reflected the preferential bioconcentration of
    the relatively toxic congeners 2,4,5,3',4'-penta-, 2,3,4,3',4'-penta-,
    and 2,3,4,5,3',4'-hexachlorobiphenyl (Parkinson, et al., 1980b). When
    Gyorkos et al. (1985) repeated the studies, the reconstituted mixture
    was inactive at the lowest dose level but exhibited mixed-type
    microsomal enzyme induction characteristics at the higher dose levels.
    Increases in the activities of several hepatic microsomal
    monooxygenases, including dimethylaminoantipyrine  N-demethylase,
    aldrin epoxidase, benzo (a)pyrene hydroxylase and
    ethoxyresorufin- O-deethylase were found.  Effects of individual congeners on liver enzymes

    The enzyme-inducing potencies of individual PCB congeners have been
    studied extensively and reviews have been published (Goldstein, 1980;
    Safe, 1984; Safe et al., 1985b), in which the following
    structure-activity relationships are proposed. The most active
    congeners with respect to the induction of aryl hydrocarbon
    hydroxylase (and toxicity), 3,4,5,4'-tetrachloro-, 3,4,3',4'-
    tetrachloro-, 3,4,5,3',4'-pentachloro-, and 3,4,5,3',4',5'-
    hexachlorobiphenyl, are substituted at both  para positions, at 2 or
    more  meta positions, but not at  ortho positions. These congeners
    can assume coplanar conformations and are approximate stereoisomers of
    2,3,7,8-tetrachlorodibenzo- para-dioxin. They resemble
    3-methylcholanthrene and 2,3,7,8-tetrachlorodibenzo- para-dioxin in
    their mode of hepatic enzyme induction, inducing hepatic microsomal
    benzo (a)pyrene hydroxylase, ethoxyresorufin- O-deethylase, and the
    cytochromes P-450 a, P-450 c, and P-450 d.

    These congeners are only present as trace compounds in commercial PCB
    mixtures, but appear in significant quantities in breast milk (Noren
    et al., 1990).

    The least active of these 4 coplanar congeners, 3,4,5,4'-
    tetrachlorobiphenyl, also shows a phenobarbital type of hepatic
    microsomal enzyme induction, inducing dimethylaminoantipyrine,
    ethylmorphine and related  N-dealkylases, biphenyl-4-hydroxylase,
    aldrin epoxidase, several  O-dealkylases, and the cytochromes P-450 a,
    P-450 b, and P-450 e. This "mixed-type" induction pattern is also
    shown by 3,4,4'-trichlorobiphenyl, and by all the mono- ortho, and at
    least 7 di- ortho, substituted analogues of the coplanar PCB
    congeners. Several of these congeners, e.g., 2,4,5,3',4'-penta,
    2,3,3',4,4'-penta, 2,3,4,5,3',4'-hexa-, and 2,3,4,5,2',4'-
    hexachlorobiphenyl are components of commercial PCB mixtures and have
    been identified in breast milk.

    Studies have revealed that 4,4'-dichlorobiphenyl, with no  meta
    substituents exhibits a PB-type induction pattern in rats. Adding
     meta substituents as in 3,4,4'-tri-, and 3,4,5,4'-tetrachloro-
    biphenyl will give a mixed PB- and 3-MC-type induction pattern. While
    3,4,3',4'-tetrachlorobiphenyl is a potent inducer of microsomal
    hepatic aryl hydrocarbon hydroxylase (AHH), it did not significantly
    increase the activities of benzo (a)pyrene-hydroxylase, [3H]-4-
    chlorobiphenyl-hydroxylase, and ethoxyresofurin- O-deethylase (EROD)
    at a dose level of 10 mol/kg (Andres et al., 1983).

    Most other PCB congeners are phenobarbital-type inducers or are
    inactive. In general, the more highly chlorinated of these congeners
    are more active inducers than the lower chlorinated ones, probably
    reflecting the relative half-lives of these compounds. The
    non-availability of 2 adjacent unhalogenated carbon atoms and
     para-substitution are 2 factors that decrease the degradability of
    the congeners and increase their inducing activity.

    Various congeners were tested for their inducing activity in
    responsive C57BL/6J mice, i.e., mice containing the Ah receptor
    protein, and in non-responsive DBA/2J mice, lacking this receptor,
    after single intraperitoneal (Robertson et al., 1984; Silkworth et
    al., 1984) or oral (Kohli et al., 1980) doses in corn oil and
    cottonseed oil, respectively. The coplanar PCBs and their  mono-ortho
    substituted analogues all induced benzo (a)pyrene hydroxylase or
    ethoxyresorufin deethylase in responsive mice, but not, or only to a
    minor degree, in non-responsive mice. Most tested  mono-ortho
    substituted analogues of coplanar PCBs slightly induced aminopyrine
     N-demethylase in both strains, while the coplanar congeners did not
    induce this enzyme.

    The above structure-activity relationships were confirmed in a few
    limited studies on monkeys. Hepatic aryl hydrocarbon hydroxylase was
    induced in 3 young, male Rhesus monkeys after a single oral dose of
    1 mg of 3,4,3',4'-tetrachlorobiphenyl/kg body weight and in 3 young
    male monkeys during continued feeding of a diet containing 0.5 mg of
    3,4,5,3',4',5'-hexachlorobiphenyl/kg (McNulty, 1985). A single oral
    dose of 18 mg of 2,5,2',5',-tetrachlorobiphenyl/kg body weight,
    administered to male Rhesus monkeys, in corn oil, elevated hepatic
    cytochrome P-450 levels, while no change was noted in the activities
    of several microsomal enzymes (Allen et al., 1975b). Mono- ortho or
    di- ortho substituted analogues of coplanar PCB congeners have not
    been tested in monkeys.

    Vodicnik et al. (1980) studied the effect of 2,4,5,2',4',5'-
    hexa-chlorobiphenyl on hepatic microsomal monooxygenase activity in
    virgin or pregnant and lactating Sprague-Dawley mice and their
    offspring. A single intraperitoneal dose of 100 mg/kg body weight, was
    administered 14 days before mating. Prior to, and during early
    pregnancy, hepatic monooxygenase activity in pretreated mice was
    greater than that in the controls. No differences were found between
    pregnant and virgin mice. Mothers, pretreated with the
    hexachlorobiphenyl and sacrificed on the day of birth, had lower
    microsomal monooxygenase activity and cytochrome P-450 content than
    PCB-pretreated virgins sacrificed concurrently. No differences were
    noted between these groups of animals during lactation. Hepatic enzyme
    activities and cytochrome P-450 content were not different between
    newborn offspring of corn oil- and PCB-pretreated mothers. However,
    these parameters were elevated in 5- to 20-day postpartum nursing
    offspring from pretreated mothers, compared with those from corn
    oil-pretreated mothers suggesting the transfer of hexachlorobiphenyl
    through the breast milk in quantities sufficient to affect hepatic
    microsomal monooxygenase activity.

    In a study on ICR mice, Vodicnik (1986) administered 150 mg
    14C-2,4,2',4'-tetrachlorobiphenyl intraperitoneally, and compared
    hepatic microsomal ethoxycoumarin- O-deethylase activity and liver
    concentrations of 14C-activity. It was shown that pregnant mice were
    less responsive to the inducing effects of the tetrachlorobiphenyl
    than virgin or postpartum mice. This diminution in response may be, in
    part, responsible for the lack of elimination of the
    tetrachlorobiphenyl equivalents from the late pregnant animal during
    the 4-day experimental period (see section 6.4.3).

    Hardwick et al. (1985) studied both the time course and dose-response
    for the induction of the 2 isoenzymes and their respective mRNAs after
    administration of 3,4,5,3',4',5'-hexachlorobiphenyl to rats. It was
    concluded that the congener under study induced 2 major 3-MC-inducible
    isoenzymes of cytochrome P-450 and their mRNAs in a coordinated
    manner, probably via a common mechanism. The data are consistent with
    the hypothesis that both genes are probably regulated by a single
    receptor in the rat. The magnitude of the increase in the isoenzymes
    was greater than the increase in the amount of translationally active
    mRNA in polysomes, suggesting that other factors may also influence
    the relative induction of these P-450 isoenzymes.

    In this study, the BP-type inducers, 2,4,6,2',4',6'- and
    2,4,5,2',4',5'-hexachlorobiphenyl, both of which have been reported to
    increase the hepatic cytosolic receptor level and consequently enhance
    enzyme induction  in vivo, were not able to enhance EROD or AHH
    induction by 3,4,5,3',4'-pentachlorobiphenyl  in vitro.

    The mixed-type inducer 2,3,4,2',4',5'-hexachlorobiphenyl inhibited
    enzyme induction by 3,4,5,3',4'-pentachlorobiphenyl  in vitro, when
    used in concentrations of at least 400 times higher than that of the
    3,4,5,3',4'-pentachlorobiphenyl. Enzyme induction by 3,4,3',4'-
    tetrachlorobiphenyl was inhibited by 2,3,4,2',4',5'-hexachlorobiphenyl
    at concentrations at least 40 times higher, and enzyme induction by
    3,4,5,3',4',5'-hexachlorobiphenyl was inhibited by 2,3,4,2',4',5'-
    hexachlorobiphenyl at concentrations at least 8 times higher. Since
    the concentration of 2,3,4,2',4',5'-hexachlorobiphenyl found in human
    adipose tissue is about 300 times higher than those of the 3 coplanar
    PCBs, this inhibition of enzyme induction probably occurs after
    natural exposure to PCB mixtures. If enzyme induction by
    3,4,3',4'-tetrachlorobiphenyl and 3,4,5,3',4',5'-hexachlorobiphenyl is
    also inhibited by various concentrations of 2,3,4,2',3',4'-
    hexachlorobiphenyl, inhibition of enzyme induction after natural
    exposure to a mixture of PCBs is to be expected.

    It was concluded that the  in vitro enzyme induction by mixtures of
    PCBs cannot be determined by the simple addition of the induction by
    the individual PCBs. Possibly,  in vivo enzyme induction seems
    additive, because some compounds increase the receptor level, while
    other compounds inhibit enzyme induction (van Vliet, 1990).  Appraisal

    The liver is the organ most often implicated in the toxicity of PCBs
    in animals. Hepatotoxicity has been observed in numerous studies with
    exposed mice, rats, guinea-pigs, rabbits, dogs, and monkeys. The
    effects, which appear to be reversible at low doses, are similar among
    the species and include enzyme induction, liver enlargement, fat
    deposition, and necrosis. Enzyme induction is the most sensitive
    indicator of hepatic effects, but few studies have been designed to
    define the minimum effective doses of PCB mixtures. The liver
    enlargement is associated with hepatocyte enlargement and an increase
    in smooth endoplasmic reticulum and/or increased enzymatic activity.
    Proliferative lesions in the liver have been attributed to Aroclor
    treatment. The hepatic effects of Aroclors in animals appear to be
    typical of chlorinated hydrocarbons. Histologically-documented liver
    damage is a consistent finding among PCB-exposed animals.

    8.6.3   Effects on vitamins and mineral metabolism  Effects of PCB mixtures

    PCB mixtures have been found to decrease levels of retinol (Vitamin A)
    in the liver of rats (Innami et al., 1976; Kato et al., 1978; Hudecova
    et al., 1979), rabbits (Villeneuve et al., 1971a), and in the plasma
    of pigs (Guoth et al., 1984). Levels of thiamine (Vitamin B1) were

    decreased in the blood, liver, and sciatic nerve of rats (Yagi et al.,
    1979) and levels of pyridoxal phosphate (Vitamin B6) were decreased in
    several tissues of rats, while riboflavin (Vitamin B2) levels remained
    unaffected (Fujiwara & Kuriyama, 1977). The changes in the levels of
    retinol and thiamine were thought to be secondary to the induction of
    metabolizing enzymes (Yagi et al., 1979; Saito et al., 1982). The
    induction of these enzymes was also found to be responsible for the
    increased  de novo synthesis of L-ascorbic acid that was observed in
    the plasma, tissues, and urine of PCB-treated rats (Fujiwara &
    Kuriyama, 1977; Chakraborty et al., 1978; Chow et al., 1979; Saito et
    al., 1983). Lipid peroxidation was increased in the liver of
    PCB-treated rats and Saito et al. (1983) suggested that ascorbic acid
    may have initiated the peroxidation.

    It was shown that induction of NADP-cytochrome c reductase (EC and insufficiency of lipid peroxide scavengers, such as
    alpha-tocopherol (Vitamin E) and glutathione peroxidase (EC,
    could also be involved in the enhancement of lipid peroxidation by
    PCBs (Saito et al., 1982, 1983; Kamohara et al., 1984).

    PCB mixtures decreased the activity of both sodium/potassium- and
    magnesium-dependent adenosinetriphosphatase (EC in the
    tissues of rats (Narbonne et al., 1978; La Rocca & Carlson, 1979). The
    results of  in vitro studies on isolated rat mitochondria showed that
    PCB mixtures may act as inhibitors of respiration and uncouplers of
    oxidative phosphorylation (Sivalingan et al., 1973; Nishihara, 1983,
    1985). However, contradictory results have been obtained  in vivo
    with respect to the NAD/NADH ratio, the ADP/O ratio, and state 3 and
    state 4 respiration rates (Mehlman et al., 1974; Chesney & Allen,
    1974; Garthoff et al., 1977).

    Byrne & Sepkovic (1987) studied the  in vitro incorporation of
    monovalent cations into rat erythrocytes as a model for evaluating the
    impairment of electrogenic transport by PCBs. Female, Sprague-Dawley
    rats were fed 50 mg Aroclor 1242 or 1254/kg diet for 7 months. The
    uptake of 86Rb by erythrocytes in the Aroclor 1254 group was
    depressed compared with that in the control group in K+ - depleted
    culture media. No changes were observed with Aroclor 1242. A reduction
    in 86Rb incorporation was also seen in erythrocytes from the Aroclor
    1254 group in a Na+ -depleted medium. Ouabain did not have any
    effect in the Aroclor 1254 group, because Aroclor 1254 suppressed the
    cationic transport maximally. This study provides evidence that PCBs
    (Aroclor 1254) can damage the cell sufficiently to decrease the active
    transport of monovalent cations.

    Male Fischer 344 rats were dosed daily, intragastrically, for 5, 10,
    or 15 weeks with 0, 0.1, 1, 10, or 25 mg Aroclor 1254/kg body weight
    in corn oil, to investigate the effects on calcium metabolism, femur
    morphometry, and nephrotoxicity. The relative liver weights were
    increased significantly with doses of 1.0 mg/kg or more after 5 weeks
    treatment. The relative kidney weights were increased after 15 weeks
    treatment in the 10 and 25 mg/kg groups. Hypercalcaemia was present in
    the 25 mg/kg group after 5 and 10 weeks treatment, but not after 15
    weeks. Serum triglyceride levels were elevated after 5 weeks
    treatment, but decreased after 10 and 15 weeks. Serum cholesterol
    levels were increased at the 2 higher dose levels with all 3 lengths
    of treatment. Urinary alkaline phosphatase and lactate dehydrogenase
    activities were elevated at 5, 10, and 15 weeks of treatment. Femur
    density was increased at the 10 mg/kg dose level after 5 weeks, and at
    all dose levels after 10 and 15 weeks. Cross-sectional, medullary, and
    cortical areas of the midpoint of the femur were significantly
    decreased at the higher dose levels after 10 and 15 weeks of exposure.
    The per cent medullary area was decreased after 10 and 15 weeks
    treatment indicating a decrease in medullary size and also a decrease
    relative to the cortical bone area. The result was weaker bones after
    15 weeks at the highest dose level. Thus, PCB exposure affects calcium
    metabolism and bone morphometry (Andrews, 1989).  Effects of individual congeners

    3,4,3',4'-Tetrachlorobiphenyl induced a decrease in serum and liver
    retinol and retinyl palmitate in C57BL/Rij mice. In "non-responsive"
    DBA/2 mice, only serum retinol was decreased. The time and
    dose-responses observed suggested that the difference in aryl
    hydrocarbon hydroxylase responsiveness was not directly involved in
    the effects on retinoid levels (Brouwer et al., 1985).

    Powers et al. (1987) administered female, Sprague-Dawley rats single,
    intraperitoneal injections of 1, 5, or 15 mg 3,4,3',4'-
    tetrachlorobiphenyl/kg body weight and found a dose-related
    depression of plasma retinol levels, 24 h after treatment. The loss of
    plasma retinol appeared to be a function of depressed levels of the
    retinol-binding protein (RBP)-transthyretin ternary complex. No free
    retinol was observed in the plasma. Hepatic retinyl palmitate
    hydrolase (RPH) activity was depressed and highly and positively
    correlated with the plasma retinol levels. Doses of either
    2,4,5,2',4',5'- and 3,4,5,3',4',5'-hexachlorobiphenyl, equimolar to
    the 15 mg/kg tetrachlorobiphenyl dose, failed to cause a similar
    depression in plasma retinol in treated female rats.

    A study was carried out to investigate the effects of PCBs on retinoid
    homeostasis in Sprague-Dawley rats. Female Sprague-Dawley/Rij rats
    were fed a Vitamin A-deficient diet for 12-16 weeks. Serum retinol
    concentrations at the end of this period were decreased to
    approximately 10% of the normal retinol level. The rats were repleted
    with radiolabelled [3H]retinol by feeding a diet containing 18.5 MBq
    (8000 IU) of retinol/kg diet for 14 days. Saturation in the blood was
    reached after 6 days [3H]retinol repletion. On day 7, the rats were
    either treated with an intraperitoneal dose of 3,4,3',4'-
    tetrachlorobiphenyl (15 mg/kg) in corn oil, or corn oil alone.
    Exposure to tetrachlorobiphenyl resulted in significant reductions in
    both retinol and retinyl ester concentrations in the liver and lung to
    25% and 44% of the controls, respectively, and a reduction of retinol
    in the heart of 35% of the controls. No changes in concentrations were
    observed in the skin and kidneys (Brouwer et al., 1988).

    Female WAG/Rij rats received a single ip injection of corn oil, or 15
    or 200 mg 3,4,3',4'-tetrachlorobiphenyl/kg body weight and were killed
    on days 1,3,7, or 14 to study the effects on serum and hepatic
    retinoid contents and liver morphology. There was a significant
    increase in liver weight at the highest dose level after 3, 7, and 14
    days. There was a rapid increase in the 3H-tetrachlorobiphenyl levels
    present after 7 days, after which a rapid decline occurred.
    Tetrachlorobiphenyl induced a significant decrease in serum retinol
    content in the 200 mg tetrachlorobiphenyl group on days 3 and 7. The
    same was found for the retinol and retinyl palmitate contents of the
    liver. Ultrastructural alterations in the hepatocytes, such as
    proliferation and vesiculation of the endoplasmic reticulum and
    mitochondrial enlargement with inclusions, were found (Durham &
    Brouwer, 1989).

    2,2',5,5'-Tetrachlorobiphenyl caused inhibition of Ca/Mg- and
    Mg-dependent adenosinetriphosphatase in the liver of rats (Lin et al.,
    1979).  In vitro, several PCB congeners inhibited Na/K- and
    Mg-dependent adenosinetriphosphatase. Although a general trend towards
    increased inhibition, paralleling increased chlorination, was
    observed, no correlation was evident between chlorine substitution
    patterns and inhibitory activity (La Rocca & Carlson, 1979).

    8.6.4   Effects on the gastrointestinal tract

    Effects on the stomach have been studied or observed by Allen &
    Norback (1973); Allen et al. (1974a); Allen (1975); Becker et al.
    (1979) and Tryphonas et al. (1986a) in monkeys. Oral administration of
    Aroclor 1242, 1248, or 1254 to monkeys produced gastritis, which
    progressed to hypertrophy and hyperplasia of the gastric mucosa.
    Related effects included mucous-filled cysts that penetrated the
    muscularis mucosa. These effects were initiated by exposure as low
    and/or short as a single gavage dose of 1.5 g Aroclor 1248/kg body
    weight, 25 mg Aroclor 1248/kg diet for up to 1 year, 3 mg of Aroclor
    1242/kg diet, for 71 days, or 280 g/kg body weight for 28 months.

    In studies on monkeys, Becker et al. (1979) carried out stomach
    biopsies and found microscopically apparent arrest of the
    differentiation of generative cells of the isthmus and neck into
    parietal and zymogenic cells. Mature parietal and zymogenic cells,
    which were found only in the bases of the glands, showed signs of
    injury, such as dilatation of the rough endoplasmic reticulum on the
    zymogenic cells, irregularity of the mitochondria and irregular
    luminal membranes in parietal cells, and an increase in the number of
    autophagic vesicles on both types of cell (see

    The Aroclor-induced gastric lesions, which occurred mainly along the
    greater curvature of the stomach (not in the cardiac or pyloric
    regions) and did not occur in other sections of the gastrointestinal
    tract, have only been observed in pigs and monkeys (Hansen et al.,
    1976b; Becker et al., 1979; Drill et al., 1981). The gastric effects
    may therefore be species specific. Aroclor 1254 induced metaplasia and
    adenocarcinoma in the glandular stomach of F344 rats (see section

    8.6.5   Effects on lipid metabolism  Effects of PCB mixtures

    Consistent with the histopathological observation of fatty
    degeneration in the liver (see section 8.2.1), short-term exposure to
    commercial mixtures of PCBs induced increases in the contents and
    concentrations of total lipids, triglycerides, cholesterol, and/or
    phospholipids in this organ of the rat and rabbit (Litterst et al.,
    1972; Bruckner et al., 1974; Itokawa et al., 1976; Garthoff et al.,
    1977; Hinton et al., 1978; Ishidate et al., 1978; Dzogbefia et al.,
    1978; Yagi, 1980; Kato & Yoshida, 1980, 1981; Kato et al., 1982).

    Litterst et al. (1972) exposed male Osborne-Mendel rats, for 4 weeks,
    to diets containing Aroclor 1242, 1248, 1254, or 1260 at levels of, or
    between, 0.5, 5.0, 50, and 500 mg/kg (equivalent to 0.025, 0.25, 2.5,
    and 25 mg/kg body weight). Aroclor 1248 caused the highest
    dose-related increase in the triglyceride concentration in the liver,
    which was significant at 500 mg/kg diet (equivalent to 25 mg/kg body
    weight). The lowest-observed-effect level was reported by Bruckner et
    al. (1974), who exposed male Sprague-Dawley rats to 0, 5, or 25 mg of
    Aroclor 1242/kg diet (equivalent to 0, 0.3, and 1.5 mg/kg body weight,
    respectively), for 2, 4, or 6 months and found a slight increase in
    the concentrations of total lipids in the liver at both exposure

    Levels of total lipids, triglycerides, and/or cholesterol in the serum
    of rats and rabbits, exposed to PCB mixtures, were found to be
    increased (Koller & Zinkl, 1973; Allen et al., 1976; Itokawa et al.,
    1976; Garthoff et al., 1977; Zinkl, 1977; Kato & Yoshida, 1980, 1981;
    Yagi, 1980; Kato et al., 1982; Baumann et al., 1983; Hladk et al.,
    1983; Carter, 1985). Wistar rats received Clophen A50, twice weekly,
    by gavage, at levels of 2, 10, 50, 150, or 250 mg/kg body weight, for
    6 weeks. Serum triglyceride and cholesterol levels were increased in a
    dose-related manner at 50 and 2 mg/kg body weight, respectively
    (Baumann et al., 1983). Decreased serum triglyceride levels were
    reported by Kato et al. (1982). Fischer rats exposed for 8 days to
    Aroclor 1254 in the diet at levels of 8 mg/kg (0.4 mg/kg body weight)
    or more showed a dose-related increase in serum total cholesterol
    concentrations. Hypercholesterolaemia was not found at 4 mg/kg diet
    (Carter, 1985). Studies on monkeys exposed to Aroclors 1248 or 1254
    for 1-2 years revealed lowered serum levels of total lipids,
    triglycerides, and cholesterol (Barsotti et al., 1976; Arnold et al.,
    1984). The cause of these changes may be an altered synthesis and/or
    lipoprotein transport in the liver. No increase in the rate of
    synthesis of liver triglycerides was found following intraperitoneal
    exposure of rats to 3-8 daily doses of 50 mg of Aroclor 1254/kg body
    weight (Hinton et al., 1978; Sandberg & Glaumann, 1980). The observed
    increases in the half-lives of liver triglycerides and phospholipids
    (Hinton et al., 1978) and the observed increase in the number of Very
    Low Density Lipoproteins (VLDL) in the liver, without a change in
    lipid composition (Sandberg & Glaumann, 1980), seems to be indicative
    of impaired transport of these lipids from the liver to the blood.
    This was also demonstrated by the repression in serum VLDL and in the
    incorporation of tritiated water in serum total lipids following
    tritiated water injection, found in rats exposed for 24 days to a
    low-protein diet containing 1000 mg of Aroclor 1248/kg (50 mg/kg body
    weight) (Kato et al., 1982). Sandberg & Glaumann (1980) observed an
    impaired transport of VLDL from the endoplasmic reticulum to the Golgi
    apparatus. This compares well with the observed flattening of the
    Golgi apparatus, which also lacks secretory vesicles with lipoprotein
    particles (Hinton et al., 1978). No explanation was found in the
    available literature for the observed increase in serum triglyceride
    levels in rats.

    Ishidate et al. (1978) measured a decreased rate of synthesis of
    phospholipids, especially of phosphatidyl choline, in the liver of
    rats that had received 2 daily doses of a PCB mixture at 100 mg/kg
    body weight, composed mainly of tetrachlorobiphenyl isomers. Decreased
    phospholipid synthesis was also observed by Hinton et al. (1978). The
    accumulation of phospholipids in the proliferated endoplasmic
    reticulum was ascribed to the observed depression of the secretion of
    lipoproteins into blood (Hinton et al., 1978; Ishidate et al., 1978;
    Sandberg & Glaumann, 1980) and to a depressed catabolism of liver
    phospholipids (Ishidate et al., 1978).

    The synthesis of cholesterol in the rat liver may be increased by PCB
    mixtures considering the increased concentration of labelled
    cholesterol in the liver following 3H20-injection (Kato et al.,
    1982) or 14C-glucose or 14C-acetate administration (Yagi, 1980) in
    rats that had been exposed for 3-5 weeks to diets containing 1000 mg
    of Aroclor 1248/kg diet or 500 mg Kanechlor 500/kg diet, respectively.
    It was also shown that the activity of 3-hydroxy-3-methylglutaryl
    Coenzyme A reductase (EC was increased in rats following a
    6-day exposure to a diet containing 1000 mg Aroclor 1248/kg diet
    (equivalent to 50 mg/kg body weight) (Kato & Yoshida, 1980). The
    enhanced synthesis of cholesterol has to compete with an enhanced
    degradation, as Aroclor 1248 has been shown to induce cholesterol
    7-alpha-hydroxylase (EC in rats (Quazi et al., 1984).
    Decreased biosynthesis of liver cholesterol was found in rats exposed
    for 30 days to Aroclor 1254 at a dietary level of 500 mg/kg (Kling &
    Gamble, 1982). Hypercholesterolaemia in PCB-exposed rats can be
    explained partly by an increased synthesis of cholesterol and/or an
    increase in serum high density lipoprotein cholesterol, which was
    observed in several studies (Ishikawa et al., 1978; Yagi, 1980; Kato &
    Yoshida, 1981; Carter, 1985).

    Isolated hepatocytes were capable of secreting protein and
    triacylglycerol in the form of VLDL into serum-free media. Eighty per
    cent of 2,4,5,2',4',5'-hexachlorobiphenyl released from hepatocytes
    was in association with VLDL, the remainder being in association with
    protein (Gallenberg & Vodicnik, 1987).  Effects of individual congeners

    Charles-River CD rats received a single oral dose of 3,4,5,3',4',5'-,
    2,4,5,2',4',5'-, or 2,3,5,2',3',5'-hexachlorobiphenyl in cotton-seed
    oil. After 72 h, all isomers had increased the levels of total lipids
    in the liver. 3,4,5,3',4',5'-Hexachlorobiphenyl had the most
    pronounced effect. This isomer was the only one that increased the
    levels of total cholesterol, cholesterol esters, and triglycerides in
    the liver, while the other 2 isomers slightly increased the content of
    liver phospholipids (Kohli et al., 1979).

    Shireman (1988) studied the lipoprotein-mediated transfer of
    2,4,5,2',4',5'-hexachlorobiphenyl into cultured human fibroblasts, and
    found that the plasma lipoproteins may play a role in the distribution
    of this hexachlorobiphenyl to peripheral cells. Using normal skin
    fibroblasts incubated with medium containing serum LDL or high density
    lipoproteins (HDL) labelled with the 14C-hexachlorobiphenyl, the
    author characterized the cellular incorporation, and efflux from
    cells, of this congener and concluded that HDL might be involved in
    the delivery of hexachlorobiphenyl to cells and not, as generally
    thought, in the transport from cells.

    2,4,5,2',4',5'-Hexachlorobiphenyl was shown to be distributed among
    rat and human plasma lipoproteins and protein  in vitro. It was
    readily transferred among plasma constituents and its distribution was
    related to the triacylglycerol:protein ratio in the plasma. One h
    following intravenous administration of 70 g labelled
    hexachlorobiphenyl to virgin, female Sprague-Dawley rats, the
    hexachlorobiphenyl was primarily distributed to low density
    lipoprotein (LDL) with the hypertriglyceridemia of late pregnancy;
    more than 70% of circulating hexachlorobiphenyl was associated with
    very low density lipoproteins (VLDL). VLDL is a major substrate for
    mammary gland lipoprotein lipase, which is elevated during lactation.
    When hexachlorobiphenyl was complexed with human VLDL and injected
    intravenously into late pregnant mice, mammary gland concentrations of
    the compound exceeded those in the adipose tissue at all sacrifice
    times between 5 min and 6 h (Gallenberg & Vodicnik, 1987; Gallenberg
    et al., 1987).

    8.6.6   Effects on porphyrin metabolism  Effects of PCB mixtures

    Hepatic porphyria has been induced by a number of commercial PCB
    mixtures (Clophen A60; Phenochlor DP6; Aroclor 1016, 1232, 1242, 1254,
    and 1260; Kanechlor 400, 500, and 600) in mice, rats, rabbits,
    chickens, and Japanese quail. Young rats, guinea-pigs, and minks seem
    to be less sensitive (Strik, 1973). The porphyria was characterized by
    the presence of pigment in the liver which fluoresced red under UV
    radiation (Vos & Beems, 1971; Kimbrough et al., 1972; Vos &
    Notenboom-Ram, 1972; Zinkl, 1977; Honda et al., 1983), an increase in
    the concentration of porphyrins in the liver (Goldstein et al., 1974,
    1975; Grote et al., 1975; Iverson et al., 1975; Kawanishi et al.,
    1975) and an increase in the concentrations of delta-aminolevulinic
    acid, porphobilinogen, and porphyrins in the urine or faeces (Vos &
    Beems, 1971; Vos & Notenboom-Ram, 1972; Goldstein et al., 1974, 1975;
    Baumann et al., 1983; Honda et al., 1983). Vos & Beems (1971) found
    increased faecal elimination of coproporphyrin and protoporphyrin in
    rabbits dermally treated with 118 mg Aroclor 1260/day (free of PCDFs),
    5 days/week, for 36 days and Vos & Notenboom-Ram (1972) found the same
    results when female, New Zealand rabbits received a 120 mg application
    of Aroclor 1260 on the shaved skin, 5 days/week, for 4 weeks.

    When Sherman rats were exposed for up to 13 months to a diet
    containing 100 mg Aroclor 1254/kg or for up to 26 weeks to Aroclor
    1242 at 100 or 500 mg/kg diet (equivalent to 5 and 25 mg/kg body
    weight, respectively) a delayed onset of porphyria was noted after 2-7
    months of exposure. The porphyria was mainly characterized by the
    excretion and hepatic storage of uroporphyrin and heptacarboxy-
    porphyrin, resembling human porphyria cutanea tarda (Goldstein et al.,

    1974, 1975). A dose-dependent increase in the concentration of liver
    porphyrins was observed in female, Sprague-Dawley rats receiving 21
    daily doses of Aroclor 1242 (by gavage) in corn oil at 10 or 100 mg/kg
    body weight, but not at 1 mg/kg body weight. Female rats were more
    sensitive than male rats and Aroclor 1016 at the same dietary level
    had less effect than Aroclor 1242 (Iverson et al., 1975). Others also
    noted the greater effect of higher chlorinated PCB mixtures on liver
    and urinary levels of porphyrins (Goldstein et al., 1974, 1975;
    Kawanishi et al., 1975). Kawanishi et al. (1973, 1974) showed that
    administration, in the diet, of Kanechlors KC-300 and KC-500 to rats
    at 500 mg/kg produced a marked increase in urinary excretion of copro-
    and uroporphyrins, and in faecal elimination of protoporphyrin, but no
    increases were observed with Kanechlor KC-400.

    Increased urinary coproporphyrin levels were found in male
    Sprague-Dawley rats exposed for 2, 4, or 6 months to a diet containing
    Aroclor 1242 at 5 or 25 mg/kg (equivalent to 0.25 and 1.25 mg/kg body
    weight) (Bruckner et al., 1974).

    Porphyria in rats and rabbits has been associated with the observed
    stimulation of delta-aminolevulinate synthase (EC, the
    rate-limiting enzyme in the haem synthesis of porphyrins (Goldstein et
    al., 1974, 1975; Grote et al., 1975; Drill et al., 1981; Hill, 1985),
    and with the inhibition of uroporphyrin decarboxylase (EC,
    as measured in chick embryo cells and chicken erythrocytes  in vitro
    (Kawanishi et al., 1983; Sano et al., 1985). Seki et al. (1987)
    observed 80% inhibition of liver uroporphyrin decarboxylase together
    with a 15-fold increase in the activity of liver delta-aminolevulinate
    synthase and accumulation in the liver of a large amount of
    uroporphyrin in C57BL/6 mice exposed for 3 weeks to Kanechlor 500 at a
    dietary dose of 500 mg/kg. Liver microsomal cytochrome P-450 was
    increased and induction of microsomal enzymes was observed. The
    effects were less outstanding in ddY mice whereas liver cytosol levels
    of the PCBs were comparable in both strains. The authors postulated
    that the development of porphyria is causally related to the
    inhibition of uroporphyrin decarboxylase rather than the induction of
    drug metabolizing function. Porphyria would develop only when the
    ratio of hepatic uroporphyrin decarboxylase and delta-aminolevulinate
    synthase decreased to less than 1.0.  Effects of individual congeners

    The levels of coproporphyrin and protoporphyrin found in the faeces of
    rabbits, dermally exposed to 5 doses/week of 120 mg of
    2,4,5,2',4',5'-hexachlorobiphenyl (no dibenzofurans detected) in
    isopropanol, for 4 weeks, were more elevated than those in the faeces
    of rabbits exposed similarly to Aroclor 1260 (Vos & Notenboom-Ram,
    1972). Koss et al. (1980) also found 2,4,5,2',4',5'-hexachlorobiphenyl

    highly effective in inducing porphyria in female rats receiving,
    orally, 64 mg of this PCB-congener/kg body weight in oil, once every 2
    days, for 10 weeks. In mice receiving a diet containing 300 mg of one
    of various tetrachlorobiphenyls, hexachlorobiphenyls, or
    Kanechlors/kg, for 14 weeks, the most pronounced increases in the
    levels of coproporphyrin and protoporphyrin in the liver were found in
    animals fed 3,4,5,3',4',5'- and 2,4,6,2',4',6'-hexachlorobiphenyl, and
    Kanechlor 600, followed by animals fed 3,5,3',5'- and 2,5,2',5'-
    tetra-chlorobiphenyls and Kanechlor 500. No porphyrinogenic action was
    found in mice fed 3,4,3',4'-, 2,4,2',4'-, 2,3,2',3'-, or 2,6,2',6'-
    tetra-chlorobiphenyl, 2,3,4,2',3',4'-hexachlorobiphenyl, or Kanechlor
    400 (Kawanishi et al., 1975). Accumulation of uroporphyrins was
    observed in the livers of "responsive" C57BL/6 mice treated with
    3,4,5,3',4',5'-hexachlorobiphenyl, but not in the livers of
    "non-responsive" ddY mice. It was suggested that induction of
    apocytochrome P-450 may take part in inducing porphyrin synthesis
    (Sano et al., 1985).

    Sano et al. (1985) studied the mechanism of the porphyrinogenic
    activity of PCBs using cultured chick embryo liver cells to examine
    the relationship between the induction of delta-aminolaevulinic acid
    (ALA) synthetase and the inhibition of uroporphyrinogen dicarboxylase.
    The porphyrinogenic effect of PCBs exhibited a defined
    structure-activity relationship in that only 3,4,3',4'-tetrachloro-
    and 3,4,5,3',4',5'-hexachlorobiphenyl out of 9 biphenyls produced a
    marked accumulation of uroporphyrin in the liver cells. In
    ALA-supplemented cultures, these 2 congeners led to the accumulation
    of a large amount of uroporphyrin III, whereas with the other PCBs
    (which were weak inducers of porphyrin synthesis) the accumulated
    porphyrin was mostly protoporphyrin. These results suggested that the
    active inducers of porphyrin synthesis also inhibit uroporphyrinogen
    decarboxylase, in 2 steps, i.e., first, in the formation of
    hexacarboxylic porphyrinogen III from heptacarboxylic porphyrinogen
    III, and, second, in the formation of heptacarboxylic porphyrinogen
    III from uroporphyrinogen III. The inhibition of uroporphyrinogen
    decarboxylase leads to a depletion of haem. In addition, induction of
    apocytochrome P-450 by PCBs may contribute to a decrease of haem. As a
    result, synthesis of ALA synthetase increases, leading to an
    accumulation of uroporphyrin in liver.

    8.6.7   Effects on the endocrine system  Effects of PCB mixtures

    The underlying cause of the reproductive toxicity of PCBs, described
    in section 8.4, may be alterations in hormonal receptor binding and/or
    alterations in the steroid hormone balance through effects on
    metabolism and excretion.

    Precocious vaginal opening was observed in neonatal Sprague-Dawley
    rats receiving a subcutaneous dose of 10 mg of Aroclor 1221
    (2000 mg/kg body weight) in sesame oil on days 2 and 3 of life. At 6
    months of age, these females showed persistent vaginal estrus and
    anovulation, despite no further exposure to Aroclor 1221. Doses of
    Aroclor 1221, 1242, 1254, or 1260 at 1 mg/kg were without effect.
    Groups of 22-day-old Sprague-Dawley rats were injected subcutaneously
    with Aroclor 1221 or 1242 at 1, 10, 100, or 1000 mg/kg body weight or
    Aroclor 1254 or 1260 mixed in sesame oil at 1, 10, or 100 mg/kg body
    weight. Uteri were weighed. New-born female pups were injected
    subcutaneously on the second and third day postpartum with Aroclor
    1221 at 1 or 10 mg/day or Aroclor 1242, 1254, or 1260 in sesame oil or
    dimethylsulfoxide at 1 mg/day. Pups were weaned at 21 days and
    examined daily from the 25th day of puberty. Animals were sacrificed
    at 7 or 8 months, at which time organs were examined. A significant
    uterotrophic response was noted with 1000 mg Aroclor 1221/kg, but not
    with the other PCBs (Gellert, 1978).

    Indirect evidence for a weak estrogenic activity of PCBs was found for
    various Aroclors by the glycogen response of immature rat uterus
    (Bitman & Cecil, 1970; Bitman et al., 1972; Ecobichon & Mackenzie,
    1974) or the less sensitive uterotropic response, observed in immature
    rats exposed to Aroclor 1221, 1232, or 1248, but not in immature rats
    exposed to Aroclor 1254 or 1260 (Ecobichon & Mackenzie, 1974; Gellert,
    1978). More direct evidence is the inhibition  in vitro of the
    binding of labelled 17-beta-estradiol to the rat uterine receptor by
    Aroclors 1221 and 1254 (Nelson, 1974).

    Pregnant mares' serum was administered to immature female outbred rats
    on day 29 postpartum, and, 60 h later, rats were injected with human
    chorionic gonadotrophin. On day 34, the animals were divided into
    groups and treated orally with sesame oil (control), or 20 mg PCBs/kg
    (Clophen A30). Two days later they were killed and the ovaries removed
    and analysed for  in vitro synthesis of progesteron (unincubated,
    incubated, and incubated with luteinizing hormone). The addition of
    luteinizing hormone resulted in an approximately 100% increase in
    progesterone synthesis above basal level with tissue exposed to PCBs.
    With control tissue there was a 31% increase with luteinizing hormone
    (Fuller et al., 1980).

    PCBs induced a decrease in gonadal steroid hormone levels in rats,
    minks, seals, and monkeys. When, after confirmed ovulation, mature
    female Rhesus monkeys were exposed during the following cycle to daily
    gavage doses of 4, 16, or 64 mg of Clophen A30/kg body weight, for 28
    days, ovulation was blocked in 2 out of 4 treated monkeys. One out of
    16 controls was anovulatory. The levels of luteinizing hormone and
    follicle-stimulating hormone were not changed by the treatment (Muller
    et al., 1978).

    Plasma progesterone levels were decreased in female rats exposed for
    36 weeks to a dietary level of Aroclor 1242 of 75 mg/kg (equivalent to
    3.7 mg/kg body weight) (Jonsson et al., 1976), and in female minks
    exposed for 12.5-14.5 months to a dietary level of 2.5 mg Aroclor
    1254/kg (equivalent to 0.25 mg/kg body weight) (Aulerich et al.,
    1985). The decreased levels of gonadal hormones can be explained by
    enhanced metabolism of steroids, which are normal substrates for
    microsomal enzymes. Increases in the formation of the metabolites of
    progesterone and/or testosterone were measured in rats
    intraperitoneally exposed 1-5 times to Aroclor 1260, Aroclor 1254, or
    Kanechlor 400 (Krogh Derr, 1978; Lin et al., 1982; Yoshihara et al.,
    1982). In contrast with these findings, increased testosterone levels
    were found in male piglets exposed for 6-12 weeks to Aroclor 1232,
    1242, or 1254 at 250 mg/diet. This increased production of
    testosterone was related to increased relative testes weights
    (Platonow et al., 1976).

    Female rhesus monkeys  (Macaca mulatta) were administered gelatin
    capsules containing daily doses of 0, 5, 20, 40, or 80 g Aroclor
    1254/kg body weight, dissolved in corn oil plus glycerol. After
    approximately 2 years of dosing, when the monkeys were considered to
    be in a state approaching adipose-tissue PCBs equilibrium, each dose
    group of 16 animals was divided into 2 test groups. Daily blood
    samples from both test groups were acquired for estrogen and
    progesteron analysis during one menstrual cycle. Serum estrogen and
    progesteron concentrations in PCB-dosed monkeys were comparable with
    those in the controls, except the luteal phase progesterone levels in
    monkeys dosed with 20 and 80 g/kg. There were no apparent
    treatment-related differences in the incidence of anovulatory cycles
    or in the temporal relationship between the estrogen peak and mensus
    onset, mensus end, or the progesterone peak. Mean PCB concentrations
    in the blood and adipose tissue for the different dose levels
    administered were as follows: blood, 1, 11, 37, 74, and 125 g/litre
    and for adipose tissue 0.79, 7.88, 22.62, 47.6 and 85.3 mg/kg tissue,
    respectively (Truelove et al., 1987).

    Effects on plasma corticosteroid levels have also been observed. The
    levels were decreased in female mice exposed to a diet containing
    25 mg Aroclor 1254/kg (equivalent to 3.7 mg/kg body weight) for 3
    weeks, and in male mice exposed to a diet containing 400 mg Aroclor
    1254/kg (equivalent to 57 mg/kg body weight) for 2 weeks. No effects
    were found on adrenal weight (Sanders & Kirkpatrick, 1975). However,
    increased levels of plasma corticosterone and enlarged adrenal glands
    were observed in male mice of another strain exposed to a diet
    containing 200 mg Aroclor 1254/kg, for 2 weeks (Sanders et al., 1977).
    Wasserman et al. (1973) reported increased plasma corticosterone
    levels in rats receiving Aroclor 1221 in the drinking-water, at a
    concentration of 250 mg/litre, for 10 weeks. This finding complies
    with morphological features of hyperfunction of the adrenal zona
    fasciculata found in rats that had received 200 mg Aroclor 1221/litre
    drinking-water, for 6 weeks.

    When female rats were exposed to Aroclor 1254, in their diet at doses
    of 0, 1, 5, 10, or 50 mg/kg (equivalent to 0, 0.05, 0.25, 0.5, and
    2.5 mg/kg body weight/day, respectively) for 5-7 months, relative
    adrenal weights as well as serum levels of corticosterone,
    dehydro-epiandrosterone, and dehydro-epiandrosterone sulfate were
    decreased in a dose- and time-related manner. In the same studies, the
    effects with less chlorinated Aroclors were less pronounced (Byrne et
    al., 1988).

    In addition, the ultrastructure of beta-cells of the pancreas of rats
    was found to be changed after 13 months of exposure to 200 mg Aroclor
    1254/litre drinking-water. These changes included marked dilatation
    and vesiculation of the rough endoplasmic reticulum, hyperplastic
    Golgi complexes with a reduction in the number of secretory granules,
    and an increase in the number of beta-acinar and acinar-beta cells.
    The changes in the pancreas were suggested to be secondary to the
    increase in the level of glucocorticoids (Wasserman et al., 1975). An
    increased relative adrenal weight was observed in pigs fed Aroclor
    1242 or 1254/kg at 20 mg/kg diet (equivalent to 0.8 mg/kg body weight)
    for 91 days (Hansen et al., 1976b).

    Thyroid hormone levels were decreased in rats exposed to Aroclor 1254
    at levels of 50 mg/kg diet or more for 4-12 weeks (Collins et al.,
    1977; Collins & Capen, 1980a). Two explanations were offered. One was
    the observed increase in biliary excretion of thyroxine and
    triiodothyroxine (Bastomsky, 1974; Collins & Capen, 1980b) and the
    larger proportion of biliary thyroxine present as glucuronide
    (Bastomsky, 1974), most likely as a result of induction of microsomal
    uridine diphosphate-glucuronosyltransferase (EC (Bastomsky &
    Murthy, 1976). The other explanation was a direct effect of PCBs on
    thyroid follicular cells. When male Holtzman or Osborne-Mendel rats
    were fed a diet containing Aroclor 1254 at a level of 0, 5, 50, or
    500 mg/kg (equivalent to 0, 0.25, 2.5, and 25 mg/kg body weight),
    thyroid follicular cells exhibited a dose-dependent hypertrophy and
    hyperplasia. An abnormal accumulation of large colloid droplets and
    irregular lysosomes in the follicular cells were observed at 5 mg/kg
    diet or more and reduced serum thyroxine occurred at 50 mg/kg diet or
    more. A no-observed-effect level could not be established. Microvilli
    were decreased in number, shortened, and irregularly branched (Collins
    et al., 1977; Kasza et al., 1978a,b; Collins & Capen, 1980a, 1980b).
    The hypothalamus-pituitary axis seems not to be affected in view of
    the observed increase in the serum level of thyroid-stimulating
    hormone and in the iodine uptake by the thyroid following PCB exposure
    (Bastomsky, 1974, 1977; Collins & Capen, 1980a).

    Collins & Capen (1980a) suggested that the well-documented
    PCBs-related disturbances in reproduction, growth, and development may
    be related to alterations in thyroid structure and function in the
    dam, fetus, or neonate. The lowering of serum thyroxine appears to be
    the combined result of a direct effect on thyroid follicular cells
    with an interference in hormone secretion plus an enhanced peripheral
    metabolism of thyroxine.  Effects of individual congeners

    Exposure of rats to various congeners produced different responses in
    steroid metabolism. The most marked effects were observed after
    exposure to 2,4,5,2',4',5'-hexachlorobiphenyl, which was found to
    decrease the half-life of progesterone (rberg & Ingvast, 1977), to
    increase hydroxylation of progesterone, testosterone, and
    androstenedione, and to decrease the 5-alpha-reduction of progesterone
    and testosterone (Dieringer et al., 1979; Yoshihara et al., 1982).
    3,4,5,3',4'-Pentachlorobiphenyl was found to depress the total
    microsomal metabolism of progesterone and testosterone, though the
    7-alpha-hydroxylation of these steroids was markedly stimulated
    (Yoshihara et al., 1982). No, or very slight, effects on steroid
    metabolism were found in rats exposed to chlorobiphenyls with 4
    chlorine atoms or less (rberg & Ingvast, 1977; Dieringer et al.,

    Yoshimura et al. (1985) described a marked induction of liver
    microsomal cytochrome P-450 and cytosolic DT-diaphorase as a cause of
    a possible disorder of steroid homeostasis and promotion of
    carcinogenicity of 4-nitroquinoline  N-oxide (4-NQO) in rats
    pretreated with 3,4,5,3',4'-pentachlorobiphenyl. The animals were
    sacrificed 5 days after pretreatment. The results of the studies
    showed that 7-alpha-hydroxylation of both progesterone and
    testosterone in liver microsomes was increased, but hydroxylation at
    the 2-alpha-, 6 alpha-, and 16 alpha-positions were depressed,
    together with 5 alpha-reduction. The induced isoenzyme P-452 was most
    responsible for the 7-alpha-hydroxylation of testosterone.

    The major component (32 mol %) of the Aroclor 1221 mixture is
    2-chlorobiphenyl. The major metabolite (4,4'-dihydroxy-2-
    chlorobiphenyl) of 2-chlorobiphenyl has been shown to have a
    significant binding activity with the soluble uterine estrogen
    receptor protein in the rat, suggesting a possible explanation for the
    unique estrogenic activity of Aroclor 1221 in the rat (Korach et al.,

    8.6.8   Immunotoxicity

    Some of the studies described below are summarized in Table 52.  Effects of PCB mixtures

    (a) Mouse

    Relative thymus and spleen weights of C57BL/6 mice were unaffected by
    exposure to Aroclor 1016, for 3-41 weeks, at a dietary level of
    167 mg/kg (Silkworth & Loose, 1979). Dietary exposure of outbred mice
     (Mus musculus) to Aroclor 1248 at 50, 100, 500, or 1000 mg/kg diet
    (equivalent to 7.1 up to 143 mg/kg body weight), for 3 or 5 weeks, did
    not elicit gross signs of immunotoxicity (Thomas & Hinsdill, 1978).

    BALB/c mice fed Aroclor 1242 at a dose-level of 0 or 167 mg/kg diet
    (equivalent to 0 and 29 mg/kg body weight) for 3-9 weeks, did not show
    adverse effects on the thymus, spleen, and lymph nodes (Loose et al.,
    1977, 1978).

    In addition, Carter & Clancy (1980) observed an increased graft versus
    host response in a decreased number of spleen cells in 4BALB/c mice,
    which were exposed to a single intraperitoneal dose of 1000 mg Aroclor
    1242/kg body weight in corn oil. Spleen enlargement and lymphocyte
    depletion were observed.

    Offspring of Swiss-Webster mice, exposed via the dams which were fed
    Aroclor 1254 at dietary levels of 10, 100, or 250 mg/kg, did not
    exhibit an altered hypersensitivity reaction to oxazoline, an altered
    anti-bovine serum albumin antibody titre or an altered degree of
    phagocytosis of sheep red blood cells by peritoneal macrophages,
    compared with controls (Talcott & Koller, 1983).

    Pathogen-free ICR/JCL mice (aged 4 weeks) were intubated orally, once
    a week, for 4 weeks, with 0, 10, or 100 g Kanechlor 500/kg body
    weight. Two days after the final treatment, half of the animals of
    each group were injected intraperitoneally with 0, 50, 250, or 500 g
     E. coli endotoxin/mouse. Sensitivity to endotoxin was determined by
    24-h mortality rate. The oral administration of Kanechlor 500 up to a
    dose level of 100 g/kg body weight did not have any effect on the
    sensitivity to the endotoxin (Oishi & Hiraga, 1980).

    The relative potencies of PCB mixtures Aroclors 1260, 1254, 1248,
    1242, 1016, and 1232 to inhibit the murine, splenic, plaque-forming
    cell response to sheep red blood cells was determined for C57Bl/6
    mice. The ED50 values for the reduction in the splenic,
    plaque-forming cells were 104, 118, 190, 391, 408, and 464 mg/kg body
    weight, respectively. It was apparent that the higher PCBs (Aroclors
    1260, 1254, and 1248) were more potent than the lower chlorinated

    Previous studies have shown that a subeffective dose of Aroclor 1254
    (25 mg/kg), interacted with an immunotoxic dose of TCDD (3.7 nmol/kg),
    resulting in, a significant antagonism of the toxicity of the latter
    compound. Co-treatment of mice with a dose of all these PCB mixtures
    at 25 mg/kg and a reconstituted PCB mixture, as occurs in breast milk,
    in combination with TCDD (3.7 nmol/kg) showed that all (except Aroclor
    1232) significantly antagonized the TCDD-mediated inhibition of the
    splenic, plaque-forming cell response in C56Bl/6 mice (Davis & Safe,

        Table 52.  The humoral and cell-mediated immunotoxicity of PCBs administered via the diet in short-term studies

    Strain          PCB-          Exposure   Parameter testedb            Resultc    LOELd       NOELd        Reference
                    mixture       period                                             (mg/kg)     (mg/kg)


    Rhesus          Aroclor       44         anti-SRBC antibody titre     D          5.0 (d)     2.5 (d)      Thomas &
                    1248                     anti-tetanus toxicoid                                            Hinsdill
                                             antibody titre               NE         -           5.0 (d)      (1978)
                                             serum gamma-globulin
                                             fraction                     D          5.0 (d)     2.5 (d)

    Cynomolgus      Kanechlor     20         anti-SRBC antibody titre     D          2 (bw)                   Hori et al.
                    400                      serum gamma-globulin                                             (1982)
                    (purified)               fraction                     D          2 (bw)

    Cynomolgus      Aroclor       21         anti-SRBC antibody titre     D          2.5 (d)                  Truelove
                    1254                     anti-tetanus toxicoid                                            et al. (1982)
                                             antibody titre               NE                     10 (d)

    Table 52.  (cont'd).

    Strain          PCB-          Exposure   Parameter testedb            Resultc    LOELd       NOELd        Reference
                    mixture       period                                             (mg/kg)     (mg/kg)


    New Zealand     Aroclor       5          anti-SRBC antibody titre     NE                     6.54 (bw)    Street &
                    1254                     serum gamma-globulin                                             Sharma
                                             levels                       NE                     6.54 (bw)    (1975)
                                             delayed hypersensitivity
                                             reaction to tuberculin       NE                     6.54 (bw)
                                             popliteal lymph node
                                             antibody-forming cells       D          0.92 (bw)   0.18 (bw)


                    Clophen       4-7        anti-tetanus toxicoid                                            Vos & van
                    A60 and                  antibody titre               D          50 (d)      10 (d)       Driel-
                    Aroclor                  delayed hypersensitivity                                         Grootenhuis
                    1260                     reaction to tuberculin       D          50 (d)      10 (d)       (1972)
                                             anti-tetanus toxicoid
                                             producing cells in
                                             popliteal lymph nodes        D          50 (d)      10 (d)

    Table 52.  (cont'd).

    Strain          PCB-          Exposure   Parameter testedb            Resultc    LOELd       NOELd        Reference
                    mixture       period                                             (mg/kg)     (mg/kg)


    Sprague-        Aroclor       1          mitogen response to                                              Bonnyns &
    Dawley          1254                     phytohaemagglutinin          I          250 (d)                  Bastomsky
                                             response to poke-weed                                            (1976)
                                             mitogen                      NE                     250 (d)
                                             serum gamma-globulin
                                             fraction                     D          250 (d)

    Sprague-        Aroclor       10         interleukin 2 production                                         Exon et al.
    Dawley          1254                     induction by KLH             D          50 (d)                   (1985)
                                             natural killer cell
                                             cytotoxicity                 D          50 (d)
                                             anti-KLH antibody titer      D          50 (d)


    ICR             Kanechlor     3          host-resistance to herpes                                        Imanishi
                    500                      simplex virus                D          33 (bw)     18 (bw)      et al. (1980)
                                             host-resistance to
                                             ectomelia virus              D          33 (bw)     18 (bw)

    Table 52.  (cont'd).

    Strain          PCB-          Exposure   Parameter testedb            Resultc    LOELd       NOELd        Reference
                    mixture       period                                             (mg/kg)     (mg/kg)

    ICR             Kanechlor     3          host-resistance to                                               Imanishi
                    500                      influenza virus              D          400 (d)     200 (d)      et al. (1984)
                                             host-resistance to
                                             Staphylococcus aureus        D          100 (d)

    ICR/JCL         Kanechlor     4          sensitivity to               NE         -           100 g/kg    Oishi &
                    500                      E. coli endotoxin            (bw)                                Hiraga

    Swiss-Webster   Aroclor       12         hypersensitivity reaction    NE         -           > 250 (d)    Talcott &
                    1254                     to oxazoline, anti-bovine                                        Koller
                                             serum albumine antibody                                          (1983)
                                             titre and phagocytosis
                                             of SRBC by macrophages

    BALB/c          Aroclor       3-6        host-resistance to                                               Loose et al.
                    1242                     endotoxin                    D          167 (d)                  (1978)
                                             host-resistance to
                                             malaria                      D          167 (d)

    BALB/c          Aroclor       6          spleen cellularity           NE                     167 (d)      Loose et al.
                    1242                     spleen PFC                   D          167 (d)                  (1977)
                                             serum immunoglobulins
                                             G1, A, M                     D          167 (d)

    Table 52.  (cont'd).

    Strain          PCB-          Exposure   Parameter testedb            Resultc    LOELd       NOELd        Reference
                    mixture       period                                             (mg/kg)     (mg/kg)

    C57BL/6         Aroclor       3-41       graft versus host response   NE                     167 (d)      Silkworth &
                    1016                     mixed lymphocyte                                                 Loose (1979)
                                             response                     I          167 (d)
                                             mitogen response to
                                             lipopolysaccharide           I          167 (d)
                                             mitogen response to
                                             concanavalin A               I          167 (d)

    Mus musculus    Aroclor       5          host-resistance to                                               Thomas &
                    1248                     Salmonella typhimurium       D          1000 (d)                 Hinsdill
                                             host-resistance to                                               (1978)
                                             endotoxin                    D          1000 (d)

    a  Up to day of primary immunization.
    b  SRBC = Sheep red blood cells; KLH = Keyhole limpet haemocyanin; PFC = Plaque forming cells.
    c  I = Increased; D = Decreased; NE = No effect found.
    d  LOEL = Lowest-observed-effect-level; NOEL = No-observed-effect level: in mg/kg diet (d)
       or mg/kg body weight per day (bw).

    A single administration of 500 mg Aroclor 1254/kg, intraperitoneally,
    inhibited the plaque-forming (PFC) response to subsequent challenge
    with sheep erythrocytes in Ah locus positive (C57Bl/6N or B6C3F1N)
    mice. However, Aroclor 1254 did not give induction in the Ah locus
    negative DBA/2N mice. When B6C3F1 mice were challenged with sheep red
    blood cells, 6 or 16 weeks after Aroclor 1524 treatment, substantial
    recovery of a PFC response was observed. In older (76-week-old) B6C3F1
    mice severe depression of the PFC response was observed.

    In contrast with its profound depression of a PFC response, Aroclor
    1254 (up to 1250 mg/kg) caused a slight increase in lymphocyte
    proliferation induced by either T or B cell mitogens. A single
    500 mg/kg dose of this Aroclor also suppressed the ability of
    recipient B6C3F1 animals to reject a challenge with either the
    syngenic fibrosarcoma (PYB6) or the gram negative pathogen  Listeria
     monocytogenes (Lubet et al., 1986).

    Heinzow et al. (1988) studied the effect of 2,4,5,2',4',5'-hexa-
    chlorobiphenyl in the E rosette formation with sheep red blood cells
    (SRBC) as one of the characteristics of human T-lymphocytes. The
    minimal concentration eliciting a significant monoclonal CD2 receptor
    antibody sparing effect was 1.5  10-10 mol/litre.

    C3H/HeN mice were treated, twice a week, for 2 or 3 weeks prior to
    mating, with olive oil or with Kanechlor 500 at an oral dose of
    50 mg/kg body weight. The offspring, some of which were nursed by
    unexposed dams, were tested for immunocompetence, 4-15 weeks after
    birth. The dams did not show any adverse effects on body weight,
    absolute spleen weight, and spleen cellularity. The B-cell activity of
    the offspring was comparable with that of controls. The helper T-cell
    activity was reduced up to 7-11 weeks after birth: the effect was more
    pronounced in prenatally-exposed groups (Takagi et al., 1989).

    (b) Rabbit

    Rabbits appear most sensitive with respect to the immunotoxicity of
    PCBs. Street & Sharma (1975) exposed groups of 5-7 New Zealand rabbits
    to Aroclor 1254 at dietary levels of 0, 3.7, 20, 45.8, or 170 mg/kg
    (equivalent to 0, 0.18, 0.92, 2.1, or 6.54 mg/kg body weight) for
    44-57 days. When compared with control animals, an increased degree of
    thymus atrophy was observed at all dose levels except for 20.0 mg/kg.
    At the 2 higher dose-levels, relative spleen weights were decreased
    and the number of germinal centres reduced. When 8 female New Zealand
    rabbits received 118 mg of Phenochlor DP6, Clophen A60, or Aroclor
    1260 (free of PCDFs), on the back skin, 5 times per week, for 38 weeks

    they showed leukopenia, thymus atrophy, and loss of germinal centres
    in the spleen and lymph nodes. No such changes were observed in the 4
    controls (Vos & Beems, 1971). Vos & Notenboom-Ram (1972) found the
    same effects in rabbits administered 120 mg Aroclor 1260 (free of
    PCDFs)/day, 5 days/week, for 4 weeks. No adverse gross immunotoxic
    effects were observed in groups of 10 New Zealand rabbits fed various
    Aroclors at dose levels of 150 mg/kg body weight, once a week, for
    12-14 weeks (Koller & Thigpen, 1973).

    When New Zealand rabbits were exposed orally, via intubation, to
    Aroclor 1242 at a dose of 150 mg/kg body weight, once a week, for 11
    weeks, the anti-pseudo rabies virus antibody titre and serum
    gamma-globulin levels were decreased (Koller & Thigpen, 1973). The
    overall picture is one of immunosuppression, though in 2 diet studies
    an increased activation of the cell-mediated immune response was
    observed (Bonnyns & Bastomsky, 1976; Silkworth & Loose, 1979).

    New Zealand rabbit offspring, exposed via the dams fed 0, 10, 100, or
    250 mg of Aroclor 1248/kg diet, showed a decrease in the delayed
    hypersensitivity reaction to dinitrofluorobenzene in the highest dose
    group. No effects were observed on the splenic, plaque-forming cell
    response, the antibody titre against sheep red blood cells, and the
    mitogen responses to Concanavalin A and Phytohaemagglutinin (Thomas &
    Hinsdill, 1980).

    (c) Guinea-pig

    Atrophy of the thymus has been reported in female guinea-pigs exposed
    to Clophen A60 or Aroclor 1260 for 4-7 weeks at a dietary level of
    50 mg/kg (equivalent to 2 mg/kg body weight). Following stimulation
    with tetanus toxoid, the authors found a lower antitoxin titre and a
    lower count of antitoxin-producing cells in comparison with control
    guinea-pigs, resulting in a significant reduction in immunoglobulins.
    The skin reaction after tuberculination in animals immunized with
    Freund's complete adjuvant (as a parameter of cell-mediated immunity)
    was also depressed at the 50 mg/kg dose level (Vos & van
    Driel-Grootenhuis, 1972).

    Female guinea-pigs fed diets containing 50 mg Aroclor 1260/kg for 6
    weeks had significantly lowered tetanus antitoxin titres, circulating
    leukocytes and lymphocytes, and thymus atrophy (Vos & van Genderen,
    1973). Also treatment with 10 mg Aroclor 1260/kg diet for 8 weeks
    produced splenic atrophy (Vos & de Roij, 1972). The Aroclor 1260 was
    free from PCDFs (no details).

    (d) Monkey

    Offspring of Rhesus monkeys appeared very sensitive to the toxic
    effects of PCB exposure during gestation and nursing, as already
    discussed in section 8.4.1. Thymic atrophy, loss of germinal centres
    and of lymph nodules of the spleen, and bone marrow hypocellularity
    were found in 3 out of 6 offspring that died during their first year
    of life due to exposure to PCBs via their mothers. The mothers, 9-12
    per group, were fed diets containing 0, 2.5, or 5.0 mg of Aroclor
    1248/kg (equivalent to 0, 0.09, and 0.2 mg/kg body weight), for 18
    months and were bred after 6 months of exposure. One mother in each
    exposed group died during exposure, showing an increased
    susceptibility to  Shigella flexneri. At autopsy, no lesions were
    observed in lymphoid organs and tissues (Allen & Barsotti, 1976;
    Barsotti et al., 1976).