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    Environmental Health Criteria 216




    DISINFECTANTS AND DISINFECTANT BY-PRODUCTS


    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 G. Amy, University of Colorado, Boulder,
    Colorado, USA; R. Bull, Battelle Pacific Northwest Laboratory,
    Richland, Washington, USA; G.F. Craun, Gunther F. Craun and
    Associates, Staunton, Virginia, USA; R.A. Pegram, US Environmental
    Protection Agency, Research Triangle Park, North Carolina, USA; and M.
    Siddiqui, University of Colorado, Boulder, Colorado, USA

    Published under the joint sponsorship of the United Nations
    Environment Programme, the International Labour Organisation and the
    World Health Organization, and produced within the framework of the
    Inter-Organization Programme for the Sound Management of Chemicals.




    World Health Organization
    Geneva, 2000

         The International Programme on Chemical Safety (IPCS),
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    WHO Library Cataloguing-in-Publication Data

    Disinfectants and disinfectant by-products.

         (Environmental health criteria ; 216)

         1.Disinfectants - chemistry     2.Disinfectants - toxicity
         3.Drinking water                4.Risk assessment
         5.Epidemiologic studies         I.Series

         ISBN 92 4 157216 7    (NLM Classification: QV 220)
         ISSN 0250-863X

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    CONTENTS

    ENVIRONMENTAL HEALTH CRITERIA FOR DISINFECTANTS AND DISINFECTANT
    BY-PRODUCTS

    PREAMBLE

    ACRONYMS AND ABBREVIATIONS

    1. SUMMARY AND EVALUATION
         1.1. Chemistry of disinfectants and disinfectant by-products
         1.2. Kinetics and metabolism in laboratory animals and humans
              1.2.1. Disinfectants
              1.2.2. Trihalomethanes
              1.2.3. Haloacetic acids
              1.2.4. Haloaldehydes and haloketones
              1.2.5. Haloacetonitriles
              1.2.6. Halogenated hydroxyfuranone derivatives
              1.2.7. Chlorite
              1.2.8. Chlorate
              1.2.9. Bromate
         1.3. Toxicology of disinfectants and disinfectant by-products
              1.3.1. Disinfectants
              1.3.2. Trihalomethanes
              1.3.3. Haloacetic acids
              1.3.4. Haloaldehydes and haloketones
              1.3.5. Haloacetonitriles
              1.3.6. Halogenated hydroxyfuranone derivatives
              1.3.7. Chlorite
              1.3.8. Chlorate
              1.3.9. Bromate
         1.4. Epidemiological studies
              1.4.1. Cardiovascular disease
              1.4.2. Cancer
              1.4.3. Adverse pregnancy outcomes
         1.5. Risk characterization
              1.5.1. Characterization of hazard and dose-response
                     1.5.1.1  Toxicological studies
                     1.5.1.2  Epidemiological studies
              1.5.2. Characterization of exposure
                     1.5.2.1  Occurrence of disinfectants and disinfectant
                              by-products
                     1.5.2.2  Uncertainties of water quality data
                     1.5.2.3  Uncertainties of epidemiological data

    2. CHEMISTRY OF DISINFECTANTS AND DISINFECTANT BY-PRODUCTS

         2.1. Background
         2.2. Physical and chemical properties of common disinfectants and
              inorganic disinfectant by-products
              2.2.1. Chlorine
              2.2.2. Chlorine dioxide

              2.2.3. Ozone
              2.2.4. Chloramines
         2.3. Analytical methods for disinfectant by-products and
              disinfectants
              2.3.1. Trihalomethanes, haloacetonitriles, chloral hydrate,
                     chloropicrin and haloacetic acids
              2.3.2. Inorganic disinfectant by-products
              2.3.3. Total organic carbon and UV absorbance at 254 nm
              2.3.4. Chloramines
         2.4. Mechanisms involved in the formation of disinfectant
              by-products
              2.4.1. Chlorine reactions
              2.4.2. Chlorine dioxide reactions
              2.4.3. Chloramine reactions
              2.4.4. Ozone reactions
         2.5. Formation of organohalogen disinfectant by-products
              2.5.1. Chlorine organohalogen by-products
              2.5.2. Chloramine organohalogen by-products
              2.5.3. Chlorine dioxide organohalogen by-products
              2.5.4. Ozone organohalogen by-products
         2.6. Formation of inorganic disinfectant by-products
              2.6.1. Chlorine inorganic by-products
              2.6.2. Chloramine inorganic by-products
              2.6.3. Chlorine dioxide inorganic by-products
              2.6.4. Ozone inorganic by-products
         2.7. Formation of non-halogenated organic disinfectant
              by-products
              2.7.1. Chlorine organic by-products
              2.7.2. Chloramine organic by-products
              2.7.3. Chlorine dioxide organic by-products
              2.7.4. Ozone organic by-products
         2.8. Influence of source water characteristics on the amount and
              type of by-products produced
              2.8.1. Effect of natural organic matter and UV absorbance 
                     at 254 nm
              2.8.2. Effect of pH
              2.8.3. Effect of bromide
              2.8.4. Effect of reaction rates
              2.8.5. Effect of temperature
              2.8.6. Effect of alkalinity
         2.9. Influence of water treatment variables on the amount and
              type of by-products produced
              2.9.1. Effect of ammonia
              2.9.2. Effect of disinfectant dose
              2.9.3. Effect of advanced oxidation processes
              2.9.4. Effect of chemical coagulation
              2.9.5. Effect of pre-ozonation
              2.9.6. Effect of biofiltration
         2.10. Comparative assessment of disinfectants
         2.11. Alternative strategies for disinfectant by-product control
              2.11.1. Source control
              2.11.2. Organohalogen by-products

              2.11.3. Inorganic by-products
              2.11.4. Organic by-products
         2.12. Models for predicting disinfectant by-product formation
              2.12.1. Factors affecting disinfectant by-product formation
                     and variables of interest in disinfectant by-product
                     modelling
              2.12.2. Empirical models for disinfectant by-product
                     formation
              2.12.3. Models for predicting disinfectant by-product
                     precursor removal
         2.13. Summary

    3. TOXICOLOGY OF DISINFECTANTS

         3.1. Chlorine and hypochlorite
              3.1.1. General toxicological properties and information on
                     dose-response in animals
              3.1.2. Reproductive and developmental toxicity
              3.1.3. Toxicity in humans
              3.1.4. Carcinogenicity and mutagenicity
              3.1.5. Comparative pharmacokinetics and metabolism
              3.1.6. Mode of action
         3.2. Chloramine
              3.2.1. General toxicological properties and information on
                     dose-response in animals
              3.2.2. Reproductive and developmental toxicity
              3.2.3. Toxicity in humans
              3.2.4. Carcinogenicity and mutagenicity
              3.2.5. Comparative pharmacokinetics and metabolism
         3.3. Chlorine dioxide
              3.3.1. General toxicological properties and information on
                     dose-response in animals
              3.3.2. Reproductive and developmental toxicity
              3.3.3. Toxicity in humans
              3.3.4. Carcinogenicity and mutagenicity
              3.3.5. Comparative pharmacokinetics and metabolism

    4. TOXICOLOGY OF DISINFECTANT BY-PRODUCTS

         4.1. Trihalomethanes
              4.1.1. Chloroform
                     4.1.1.1  General toxicological properties and
                              information on dose-response in animals
                     4.1.1.2  Toxicity in humans
                     4.1.1.3  Carcinogenicity and mutagenicity
                     4.1.1.4  Comparative pharmacokinetics and metabolism
                     4.1.1.5  Mode of action
              4.1.2. Bromodichloromethane
                     4.1.2.1  General toxicological properties and
                              information on dose-response in animals
                     4.1.2.2  Reproductive and developmental toxicity
                     4.1.2.3  Neurotoxicity
                     4.1.2.4  Toxicity in humans

                     4.1.2.5  Carcinogenicity and mutagenicity
                     4.1.2.6  Comparative phamacokinetics and metabolism
                     4.1.2.7  Mode of action
              4.1.3. Dibromochloromethane
                     4.1.3.1  General toxicological properties and
                              information on dose-response in animals
                     4.1.3.2  Reproductive and developmental toxicity
                     4.1.3.3  Neurotoxicity
                     4.1.3.4  Toxicity in humans
                     4.1.3.5  Carcinogenicity and mutagenicity
                     4.1.3.6  Comparative pharmacokinetics and metabolism
                     4.1.3.7  Mode of action
              4.1.4. Bromoform
                     4.1.4.1  General toxicological properties and
                              information on dose-response in animals
                     4.1.4.2  Reproductive and developmental toxicity
                     4.1.4.3  Neurotoxicity
                     4.1.4.4  Toxicity in humans
                     4.1.4.5  Carcinogenicity and mutagenicity
                     4.1.4.6  Comparative pharmacokinetics and metabolism
                     4.1.4.7  Mode of action
         4.2. Haloacids
              4.2.1. Dichloroacetic acid (dichloroacetate)
                     4.2.1.1  General toxicological properties and
                              information on dose-response in animals
                     4.2.1.2  Reproductive effects
                     4.2.1.3  Developmental effects
                     4.2.1.4  Neurotoxicity
                     4.2.1.5  Toxicity in humans
                     4.2.1.6  Carcinogenicity and mutagenicity
                     4.2.1.7  Comparative pharmacokinetics and metabolism
                     4.2.1.8  Mode of action
              4.2.2. Trichloroacetic acid (trichloroacetate)
                     4.2.2.1  General toxicological properties and
                              information on dose-response in animals
                     4.2.2.2  Reproductive effects
                     4.2.2.3  Developmental effects
                     4.2.2.4  Neurotoxicity
                     4.2.2.5  Toxicity in humans
                     4.2.2.6  Carcinogenicity and mutagenicity
                     4.2.2.7  Comparative pharmacokinetics and metabolism
                     4.2.2.8  Mode of action
              4.2.3. Brominated haloacetic acids
                     4.2.3.1  General toxicological properties and
                              information on dose-response in animals
                     4.2.3.2  Reproductive effects
                     4.2.3.3  Neurotoxicity
                     4.2.3.4  Toxicity in humans
                     4.2.3.5  Carcinogenicity and mutagenicity
                     4.2.3.6  Comparative pharmacokinetics and metabolism
                     4.2.3.7  Mode of action
              4.2.4. Higher molecular weight halogenated acids

         4.3. Haloaldehydes and haloketones
              4.3.1. Chloral hydrate (trichloroacetaldehyde, chloral)
                     4.3.1.1  General toxicological properties and
                              information on dose-response in animals
                     4.3.1.2  Toxicity in humans
                     4.3.1.3  Carcinogenicity and mutagenicity
                     4.3.1.4  Comparative metabolism and pharmacokinetics
                     4.3.1.5  Mode of action
              4.3.2. Halogenated aldehydes and ketones other than chloral
                     hydrate
                     4.3.2.1  General toxicological properties and
                              information on dose-response in animals
                     4.3.2.2  Toxicity in humans
                     4.3.2.3  Carcinogenicity and mutagenicity
                     4.3.2.4  Comparative pharmacokinetics and metabolism
                     4.3.2.5  Mode of action
         4.4. Haloacetonitriles
              4.4.1. General toxicological properties and information on
                     dose-response in animals and humans
              4.4.2. Reproductive and developmental toxicity
              4.4.3. Carcinogenicity and mutagenicity
              4.4.4. Comparative pharmacokinetics and metabolism
              4.4.5. Mode of action
         4.5. Halogenated hydroxyfuranone derivatives
              4.5.1. General toxicological properties and information on
                     dose-response in animals
              4.5.2. Toxicity in humans
              4.5.3. Carcinogenicity and mutagenicity
                     4.5.3.1  Studies in bacteria and mammalian cells
                               in vitro
                     4.5.3.2  Studies in experimental animals
              4.5.4. Comparative pharmacokinetics and metabolism
         4.6. Chlorite
              4.6.1. General toxicological properties and information on
                     dose-response in animals
              4.6.2. Reproductive and developmental toxicity
              4.6.3. Toxicity in humans
              4.6.4. Carcinogenicity and mutagenicity
              4.6.5. Comparative pharmacokinetics and metabolism
              4.6.6. Mode of action
         4.7. Chlorate
              4.7.1. General toxicological properties and information on
                     dose-response in animals
              4.7.2. Reproductive and developmental toxicity
              4.7.3. Toxicity in humans
              4.7.4. Carcinogenicity and mutagenicity
              4.7.5. Mode of action
         4.8. Bromate
              4.8.1. General toxicological properties and information on
                     dose-response in animals 
              4.8.2. Toxicity in humans
              4.8.3. Carcinogenicity and mutagenicity
              4.8.4. Comparative pharmacokinetics and metabolism
              4.8.5. Mode of action
         4.9. Other disinfectant by-products

    5. EPIDEMIOLOGICAL STUDIES
         5.1. Epidemiological study designs and causality of
              epidemiological associations
              5.1.1. Experimental studies
              5.1.2. Observational studies
              5.1.3. Random and systematic error
              5.1.4. Causality of an epidemiological association
         5.2. Epidemiological associations between disinfectant
              use and adverse health outcomes
              5.2.1. Epidemiological studies of cancer and disinfected
                     drinking-water
                     5.2.1.1  Cancer associations in ecological studies
                     5.2.1.2  Cancer associations in analytical studies
                     5.2.1.3  Meta-analysis of cancer studies
                     5.2.1.4  Summary of results of cancer studies
              5.2.2. Epidemiological studies of cardiovascular disease and
                     disinfected drinking-water
                     5.2.2.1  Summary of results of cardiovascular studies
              5.2.3. Epidemiological studies of adverse
                     reproductive/developmental outcomes and disinfected
                     drinking-water
                     5.2.3.1  Summary of results of
                              reproductive/developmental studies
         5.3. Epidemiological associations between disinfectant
              by-products and adverse health outcomes
              5.3.1. Epidemiological studies of cancer and disinfectant
                     by-products
                     5.3.1.1  Cancer associations in ecological studies
                     5.3.1.2  Cancer associations in analytical studies
                     5.3.1.3  Summary of results of cancer studies
              5.3.2. Epidemiological studies of cardiovascular disease and
                     disinfectant by-products
                     5.3.2.1  Summary of results of cardiovascular studies
              5.3.3. Epidemiological studies of adverse
                     reproductive/developmental outcomes and disinfectant
                     by-products
                     5.3.3.1  Summary of results of
                              reproductive/developmental studies
         5.4. Summary

    6. RISK CHARACTERIZATION

         6.1. Characterization of hazard and dose-response
              6.1.1. Toxicological studies
                     6.1.1.1  Chlorine
                     6.1.1.2  Monochloramine
                     6.1.1.3  Chlorine dioxide
                     6.1.1.4  Trihalomethanes
                     6.1.1.5  Haloacetic acids
                     6.1.1.6  Chlorate hydrate
                     6.1.1.7  Haloacetonitriles
                     6.1.1.8  MX
                     6.1.1.9  Chlorite

                     6.1.1.10 Chlorate
                     6.1.1.11 Bromate
              6.1.2. Epidemiological studies
         6.2. Characterization of exposure
              6.2.1. Occurrence of disinfectants and disinfectant
                     by-products
              6.2.2. Uncertainties of water quality data
              6.2.3. Uncertainties of epidemiological data

    7. RISK CONCLUSIONS AND COMPARISONS

         7.1. Epidemiological studies
         7.2. Toxicological studies
              7.2.1. Diversity of by-products
              7.2.2. Diversity of modes of action
              7.2.3. Reproductive, developmental and neurotoxic effects
         7.3. Risks associated with mixtures of disinfectant by-products

    8. CONCLUSIONS AND RECOMMENDATIONS

         8.1. Chemistry
         8.2. Toxicology
         8.3. Epidemiology

    9. RESEARCH NEEDS

         9.1. Chemistry of disinfectants and disinfectant by-products
         9.2. Toxicology
         9.3. Epidemiology 

    PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES

    REFERENCES

    RESUME ET EVALUATION

    RESUMEN Y EVALUACION
    

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    sent to the Chairperson and Rapporteur of the Task Group to check for
    any errors.

         It is accepted that the following criteria should initiate the
    updating of an EHC monograph: new data are available that would
    substantially change the evaluation; there is public concern for
    health or environmental effects of the agent because of greater
    exposure; an appreciable time period has elapsed since the last
    evaluation.

         All Participating Institutions are informed, through the EHC
    progress report, of the authors and institutions proposed for the
    drafting of the documents. A comprehensive file of all comments
    received on drafts of each EHC monograph is maintained and is

    available on request. The Chairpersons of Task Groups are briefed
    before each meeting on their role and responsibility in ensuring that
    these rules are followed.

    FIGURE

    WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR DISINFECTANTS AND
    DISINFECTANT BY-PRODUCTS

     Members 

    Dr G. Amy, Department of Civil, Environmental, and Architectural
         Engineering, University of Colorado, Boulder, Colorado, USA

    Mr J. Fawell, Water Research Centre, Marlow, Buckinghamshire, United
         Kingdom  (Co-Rapporteur) 

    Dr B. Havlik, Ministry of Health, National Institute of Public Health,
         Prague, Czech Republic

    Dr C. Nokes, Water Group, Institute of Environmental Science and
         Research, Christchurch, New Zealand  (Co-Rapporteur) 

    Dr E. Ohanian, Office of Water/Office of Science and Technology,
         United States Environmental Protection Agency, Washington, DC,
         USA  (Chairman) 

    Dr E. Soderlund, Department of Environmental Medicine, National
         Institute of Public Health, Torshov, Oslo

     Secretariat 

    Dr J. Bartram, Water, Sanitation and Health Unit, Division of
         Operational Support in Environment Health, World Health
         Organization, Geneva, Switzerland

    Dr R. Bull, Battelle Pacific Northwest Laboratory, Richland,
         Washington, USA

    Mr G.F. Craun, Gunther F. Craun and Associates, Staunton, Virginia,
         USA

    Dr H. Galal-Gorchev, Chevy Chase, Maryland, USA  (Secretary) 

    Mr N. Nakashima, Assessment of Risk and Methodologies,      
         International Programme on Chemical Safety, World Health
         Organization, Geneva, Switzerland

    Dr R.A. Pegram, United States Environmental Protection Agency,
         Research Triangle Park, North Carolina, USA

    Mr S.T. Yamamura, Water, Sanitation and Health Unit, Division of
         Operational Support in Environment Health, World Health
         Organization, Geneva, Switzerland

     Representatives/Observers 

    Dr N. Drouot, Dept Toxicologie Industrielle, Paris, France
         (representing European Centre for Ecotoxicology and Toxicology 
         of Chemicals)

    Mr O. Hydes, Drinking Water Inspectorate, London, United Kingdom 

    Dr B.B. Sandel, Olin Corporation, Norwalk, Connecticut, USA
         (representing American Industrial Health Council/International
         Life Sciences Institute)

    IPCS TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR DISINFECTANTS AND
    DISINFECTANT BY-PRODUCTS

         A WHO Task Group on Environmental Health Criteria for
    Disinfectants and Disinfectant By-products met in Geneva from 17 to 21
    August 1998. Dr Peter Toft, Associate Director, IPCS, welcomed the
    participants on behalf of the three IPCS cooperating organizations
    (UNEP/ILO/WHO). The Task Group reviewed and revised the draft document
    and made an evaluation of risks for human health from exposure to
    certain disinfectants and disinfectant by-products.

         The first draft of the chemistry section was prepared by G. Amy
    and M. Siddiqui, University of Colorado, Boulder, Colorado, USA; the
    toxicology section was prepared by R. Bull, Battelle Pacific Northwest
    Laboratory, Richland, Washington, USA, and R.A. Pegram, US
    Environmental Protection Agency, Research Triangle Park, North
    Carolina, USA; and the epidemiology section was prepared by G.F.
    Craun, Gunther F. Craun and Associates, Staunton, Virginia, USA.

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



                                  *  *  *



         The preparation of the first draft of this Environmental Health
    Criteria monograph was made possible by the financial support afforded
    to IPCS by the International Life Sciences Institute.

         A financial contribution from the United States Environmental
    Protection Agency for the convening of the Task Group is gratefully
    acknowledged.

    ACRONYMS AND ABBREVIATIONS

    ALAT           alanine aminotransferase
    AP             alkaline phosphatase
    ARB            atypical residual bodies
    ASAT           aspartate aminotransferase
    AWWA           American Water Works Association
    BAN            bromoacetonitrile
    BCA            bromochloroacetic acid/bromochloroacetate
    BCAN           bromochloroacetonitrile
    BDCA           bromodichloroacetic acid/bromodichloroacetate
    BDCM           bromodichloromethane
    BUN            blood urea nitrogen
    bw             body weight
    CAN            chloroacetonitrile
    CHO            Chinese hamster ovary
    CI             confidence interval
    CoA            coenzyme A
     Cmax           maximum concentration
    CMCF           3-chloro-4-(chloromethyl)-5-hydroxy-2(5H)-furanone
    2-CP           2-chloropropionate
    CPN            chloropropanone
    CT             computerized tomography
    CYP            cytochrome P450
    DBA            dibromoacetic acid/dibromoacetate
    DBAC           dibromoacetone
    DBAN           dibromoacetonitrile
    DBCM           dibromochloromethane
    DBP            disinfectant by-product
    DCA            dichloroacetic acid/dichloroacetate
    DCAN           dichloroacetonitrile
    DCPN           dichloropropanone
    DHAN           dihaloacetonitrile
    DOC            dissolved organic carbon
    ECD            electron capture detector
    ECG            electrocardiogram
    EEG            electroencephalogram
    EHEN            N-ethyl- N-hydroxyethylnitrosamine
    EPA            Environmental Protection Agency (USA)
    ESR            electron spin resonance
    FAO            Food and Agriculture Organization of the United Nations
    GAC            granular activated carbon
    GC             gas chromatography
    GGT            gamma-glutamyl transpeptidase
    GOT            glutamate-oxalate transaminase
    GPT            glutamate-pyruvate transaminase
    GSH            glutathione-SH
    GST            glutathione- S-transferase
    HAA            haloacetic acid
    HAN            haloacetonitrile
    HDL            high-density lipoprotein
    HPLC           high-performance liquid chromatography
    hprt           hypoxanthine phosphoribosyl transferase
    IARC           International Agency for Research on Cancer

    IC             ion chromatography
    i.p.           intraperitoneal
    IPCS           International Programme on Chemical Safety
    JECFA          Joint FAO/WHO Expert Committee on Food Additives
    LD50           median lethal dose
    LDH            lactate dehydrogenase
    LDL            low-density lipoprotein
    LOAEL          lowest-observed-adverse-effect level
    MA             3,4-(dichloro)-5-hydroxy-2(5H)-furanone
    MBA            monobromoacetic acid/monobromoacetate
    MCA            monochloroacetic acid/monochloroacetate
    MNU            methylnitrosourea
    MOR            mortality odds ratio
    MRI            magnetic resonance imaging
    MTBE           methyl  tert-butyl ether
    MX             3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone
    NADP           nicotinamide adenine dinucleotide phosphate
    NOAEL          no-observed-adverse-effect level
    NOEL           no-observed-effect level
    NOM            natural organic matter
    NTP            National Toxicology Program (USA)
    8-OH-dG        8-hydroxy-2-deoxyguanosine
    OR             odds ratio
    PAS            periodic acid/Schiff's reagent
    PBPK           physiologically based pharmacokinetic model
    PFBHA           O-(2,3,4,5,6-pentafluorobenzyl)-hydroxylamine
    p Ka            log acid dissociation constant
    PPAR           peroxisome proliferator activated receptor
    PPRE           peroxisome proliferator responsive element
    RR             relative risk
    SCE            sister chromatid exchange
    SD             standard deviation
    SDH            sorbitol dehydrogenase
    SE             standard error
    SGOT           serum glutamate-oxaloacetate transaminase
    SGPT           serum glutamate-pyruvate transaminase
    SMR            standardized mortality ratio
    SSB            single strand breaks
    TBA            tribromoacetic acid/tribromoacetate
    TBARS          thiobarbituric acid reactive substances
    TCA            trichloroacetic acid/trichloroacetate
    TCAN           trichloroacetonitrile
    TCPN           trichloropropanone
    TDI            tolerable daily intake
    TGF            transforming growth factor
    THM            trihalomethane
    TOC            total organic carbon
    TOX            total organic halogen
    TPA            12- O-tetradecanoylphorbol-13-acetate
    UDS            unscheduled DNA synthesis
    UV             ultraviolet
    UVA254         UV absorbance at 254 nm
     Vmax           maximum rate of metabolism
    WHO            World Health Organization

    1.  SUMMARY AND EVALUATION

         Chlorine (Cl2) has been widely used throughout the world as a
    chemical disinfectant, serving as the principal barrier to microbial
    contaminants in drinking-water. The noteworthy biocidal attributes of
    chlorine have been somewhat offset by the formation of disinfectant
    by-products (DBPs) of public health concern during the chlorination
    process. As a consequence, alternative chemical disinfectants, such as
    ozone (O3), chlorine dioxide (ClO2) and chloramines (NH2Cl,
    monochloramine), are increasingly being used; however, each has been
    shown to form its own set of DBPs. Although the microbiological
    quality of drinking-water cannot be compromised, there is a need to
    better understand the chemistry, toxicology and epidemiology of
    chemical disinfectants and their associated DBPs in order to develop a
    better understanding of the health risks (microbial and chemical)
    associated with drinking-water and to seek a balance between microbial
    and chemical risks. It is possible to decrease the chemical risk due
    to DBPs without compromising microbiological quality.

    1.1  Chemistry of disinfectants and disinfectant by-products

         The most widely used chemical disinfectants are chlorine, ozone,
    chlorine dioxide and chloramine. The physical and chemical properties
    of disinfectants and DBPs can affect their behaviour in
    drinking-water, as well as their toxicology and epidemiology. The
    chemical disinfectants discussed here are all water-soluble oxidants,
    which are produced either on-site (e.g., ozone) or off-site (e.g.,
    chlorine). They are administered as a gas (e.g., ozone) or liquid
    (e.g., hypochlorite) at typical doses of several milligrams per litre,
    either alone or in combination. The DBPs discussed here are measurable
    by gas or liquid chromatography and can be classified as organic or
    inorganic, halogenated (chlorinated or brominated) or non-halogenated,
    and volatile or non-volatile. Upon their formation, DBPs can be stable
    or unstable (e.g., decomposition by hydrolysis).

         DBPs are formed upon the reaction of chemical disinfectants with
    DBP precursors. Natural organic matter (NOM), commonly measured by
    total organic carbon (TOC), serves as the organic precursor, whereas
    bromide ion (Br-) serves as the inorganic precursor. DBP formation is
    influenced by water quality (e.g., TOC, bromide, pH, temperature,
    ammonia, carbonate alkalinity) and treatment conditions (e.g.,
    disinfectant dose, contact time, removal of NOM before the point of
    disinfectant application, prior addition of disinfectant).

         Chlorine in the form of hypochlorous acid/hypochlorite ion
    (HOCl/OCl-) reacts with bromide ion, oxidizing it to hypobromous
    acid/hypobromite ion (HOBr/OBr-). Hypochlorous acid (a more powerful
    oxidant) and hypobromous acid (a more effective halogenating agent)
    react collectively with NOM to form chlorine DBPs, including
    trihalomethanes (THMs), haloacetic acids (HAAs), haloacetonitriles
    (HANs), haloketones, chloral hydrate and chloropicrin. The dominance
    of chlorine DBP groups generally decreases in the order of THMs, HAAs
    and HANs. The relative amounts of TOC, bromide and chlorine will
    affect the species distribution of THMs (four species: chloroform,

    bromoform, bromodichloromethane [BDCM] and dibromochloromethane
    [DBCM]), HAAs (up to nine chlorinated/brominated species) and HANs
    (several chlorinated/brominated species). Generally, chlorinated THM,
    HAA and HAN species dominate over brominated species, although the
    opposite may be true in high-bromide waters. Although many specific
    chlorine DBPs have been identified, a significant percentage of the
    total organic halogens still remain unaccounted for. Another reaction
    that occurs with chlorine is the formation of chlorate (ClO3-) in
    concentrated hypochlorite solutions.

         Ozone can directly or indirectly react with bromide to form
    brominated ozone DBPs, including bromate ion (BrO3-). In the
    presence of NOM, non-halogenated organic DBPs, such as aldehydes,
    ketoacids and carboxylic acids, are formed during ozonation, with
    aldehydes (e.g., formaldehyde) being dominant. If both NOM and bromide
    are present, ozonation forms hypobromous acid, which, in turn, leads
    to the formation of brominated organohalogen compounds (e.g.,
    bromoform).

         The major chlorine dioxide DBPs include chlorite (ClO2-) and
    chlorate ions, with no direct formation of organohalogen DBPs. Unlike
    the other disinfectants, the major chlorine dioxide DBPs are derived
    from decomposition of the disinfectant as opposed to reaction with
    precursors.

         Use of chloramine as a secondary disinfectant generally leads to
    the formation of cyanogen chloride (CNCl), a nitrogenous compound, and
    significantly reduced levels of chlorine DBPs. A related issue is the
    presence of nitrite (NO2-) in chloraminated distribution systems.

         From the present knowledge of occurrence and health effects, the
    DBPs of most interest are THMs, HAAs, bromate and chlorite.

         The predominant chlorine DBP group has been shown to be THMs,
    with chloroform and BDCM as the first and second most dominant THM
    species. HAAs are the second predominant group, with dichloroacetic
    acid (DCA) and trichloroacetic acid (TCA) being the first and second
    most dominant species. 

         Conversion of bromide to bromate upon ozonation is affected by
    NOM, pH and temperature, among other factors. Levels may range from
    below detection (2 g/litre) to several tens of micrograms per litre.
    Chlorite levels are generally very predictable, ranging from about 50%
    to 70% of the chlorine dioxide dose administered.

         DBPs occur in complex mixtures that are a function of the
    chemical disinfectant used, water quality conditions and treatment
    conditions; other factors include the combination/sequential use of
    multiple disinfectants/oxidants. Moreover, the composition of these
    mixtures may change seasonally. Clearly, potential chemically related
    health effects will be a function of exposure to DBP mixtures.

         Other than chlorine DBPs (in particular THMs), there are very few
    data on the occurrence of DBPs in finished water and distribution
    systems. Based on laboratory databases, empirical models have been
    developed to predict concentrations of THMs (total THMs and THM
    species), HAAs (total HAAs and HAA species) and bromate. These models
    can be used in performance assessment to predict the impact of
    treatment changes and in exposure assessment to simulate missing or
    past data (e.g., to predict concentrations of HAAs from THM data).

         DBPs can be controlled through DBP precursor control and removal
    or modified disinfection practice. Coagulation, granular activated
    carbon, membrane filtration and ozone biofiltration can remove NOM.
    Other than through the use of membranes, there is little opportunity
    to effectively remove bromide. Source water protection and control
    represent non-treatment alternatives to precursor control. Removal of
    DBPs after formation is not viable for organic DBPs, whereas bromate
    and chlorite can be removed by activated carbon or reducing agents. It
    is expected that the optimized use of combinations of disinfectants,
    functioning as primary and secondary disinfectants, can further
    control DBPs. There is a trend towards combination/sequential use of
    disinfectants; ozone is used exclusively as a primary disinfectant,
    chloramines exclusively as a secondary disinfectant, and both chlorine
    and chlorine dioxide in either role.

    1.2  Kinetics and metabolism in laboratory animals and humans

    1.2.1  Disinfectants

         Residual disinfectants are reactive chemicals that will react
    with organic compounds found in saliva and stomach content, resulting
    in the formation of by-products. There are significant differences in
    the pharmacokinetics of 36Cl depending on whether it is obtained from
    chlorine, chloramine or chlorine dioxide.

    1.2.2  Trihalomethanes

         The THMs are absorbed, metabolized and eliminated rapidly by
    mammals after oral or inhalation exposure. Following absorption, the
    highest tissue concentrations are attained in the fat, liver and
    kidneys. Half-lives generally range from 0.5 to 3 h, and the primary
    route of elimination is via metabolism to carbon dioxide. Metabolic
    activation to reactive intermediates is required for THM toxicity, and
    the three brominated species are all metabolized more rapidly and to a
    greater extent than chloroform. The predominant route of metabolism
    for all the THMs is oxidation via cytochrome P450 (CYP) 2E1, leading
    to the formation of dihalocarbonyls (i.e., phosgene and brominated
    congeners), which can be hydrolysed to carbon dioxide or bind to
    tissue macromolecules. Secondary metabolic pathways are reductive
    dehalogenation via CYP2B1/2/2E1 (leading to free radical generation)
    and glutathione (GSH) conjugation via glutathione- S-transferase
    (GST) T1-1, which generates mutagenic intermediates. The brominated
    THMs are much more likely than chloroform to proceed through the
    secondary pathways, and GST-mediated conjugation of chloroform to GSH
    can occur only at extremely high chloroform concentrations or doses. 

    1.2.3  Haloacetic acids

         The kinetics and metabolism of the dihaloacetic and trihaloacetic
    acids differ significantly. To the extent they are metabolized, the
    principal reactions of the trihaloacetic acids occur in the microsomal
    fraction, whereas more than 90% of the dihaloacetic acid metabolism,
    principally by glutathione transferases, is observed in the cytosol.
    TCA has a biological half-life in humans of 50 h. The half-lives of
    the other trihaloacetic acids decrease significantly with bromine
    substitution, and measurable amounts of the dihaloacetic acids can be
    detected as products with brominated trihaloacetic acids. The
    half-lives of the dihaloacetic acids are very short at low doses but
    can be drastically increased as dose rates are increased.

    1.2.4  Haloaldehydes and haloketones

         Limited kinetic data are available for chloral hydrate. The two
    major metabolites of chloral hydrate are trichloroethanol and TCA.
    Trichloroethanol undergoes rapid glucuronidation, enterohepatic
    circulation, hydrolysis and oxidation to TCA. Dechlorination of
    trichloroethanol or chloral hydrate would lead to the formation of
    DCA. DCA may then be further transformed to monochloroacetate (MCA),
    glyoxalate, glycolate and oxalate, probably through a reactive
    intermediate. No information was found on the other haloaldehydes and
    haloketones. 

    1.2.5  Haloacetonitriles

         The metabolism and kinetics of HANs have not been studied.
    Qualitative data indicate that the products of metabolism include
    cyanide, formaldehyde, formyl cyanide and formyl halides.

    1.2.6  Halogenated hydroxyfuranone derivatives

         3-Chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) is the
    member of the hydroxyfuranone class that has been most extensively
    studied. From animal studies, it appears that the 14C label of MX is
    rapidly absorbed from the gastrointestinal tract and reaches systemic
    circulation. MX itself has not been measured in blood. The MX label is
    largely excreted in urine and faeces, urine being the major route of
    excretion. Very little of the initial radiolabel is retained in the
    body after 5 days.

    1.2.7  Chlorite

         The 36Cl from chlorite is rapidly absorbed. Less than half the
    dose is found in the urine as chloride, and a small proportion as
    chlorite. A significant proportion probably enters the chloride pool
    of the body, but a lack of analytical methods to characterize chlorite
    in biological samples means that no detailed information is available.

    1.2.8  Chlorate

         Chlorate behaves similarly to chlorite. The same analytical
    constraints apply.

    1.2.9  Bromate

         Bromate is rapidly absorbed and excreted, primarily in urine, as
    bromide. Bromate is detected in urine at doses of 5 mg/kg of body
    weight and above. Bromate concentrations in urine peak at about 1 h,
    and bromate is not detectable in plasma after 2 h.

    1.3  Toxicology of disinfectants and disinfectant by-products

    1.3.1  Disinfectants

         Chlorine gas, chloramine and chlorine dioxide are strong
    respiratory irritants. Sodium hypochlorite (NaOCl) is also used as
    bleach and is frequently involved in human poisoning. These exposures,
    however, are not relevant to exposures in drinking-water. There have
    been relatively few evaluations of the toxic effects of these
    disinfectants in drinking-water in experimental animals or humans.
    Evidence from these animal and human studies suggests that chlorine,
    hypochlorite solutions, chloramine and chlorine dioxide themselves
    probably do not contribute to the development of cancer or any toxic
    effects. Attention has focused on the wide variety of by-products that
    result from reactions of chlorine and other disinfectants with NOM,
    which is found in virtually all water sources.

    1.3.2  Trihalomethanes

         THMs induce cytotoxicity in the liver and kidneys of rodents
    exposed to doses of about 0.5 mmol/kg of body weight. The vehicle of
    administration significantly affects the toxicity of the THMs. The
    THMs have little reproductive and developmental toxicity, but BDCM has
    been shown to reduce sperm motility in rats consuming 39 mg/kg of body
    weight per day in drinking-water. Like chloroform, BDCM, when
    administered in corn oil, induces cancer in the liver and kidneys
    after lifetime exposures to high doses. Unlike chloroform and DBCM,
    BDCM and bromoform induce tumours of the large intestine in rats
    exposed by corn oil gavage. BDCM induces tumours at all three target
    sites and at lower doses than the other THMs. Since the publication of
    the 1994 WHO Environmental Health Criteria monograph on chloroform,
    additional studies have added to the weight of evidence indicating
    that chloroform is not a direct DNA-reactive mutagenic carcinogen. In
    contrast, the brominated THMs appear to be weak mutagens, probably as
    a result of GSH conjugation.

    1.3.3  Haloacetic acids

         The HAAs have diverse toxicological effects in laboratory
    animals. Those HAAs of most concern have carcinogenic, reproductive
    and developmental effects. Neurotoxic effects are significant at the
    high doses of DCA that are used therapeutically. Carcinogenic effects

    appear to be limited to the liver and to high doses. The bulk of the
    evidence indicates that the tumorigenic effects of DCA and TCA depend
    on modifying processes of cell division and cell death rather than
    their very weak mutagenic activities. Oxidative stress is also a
    feature of the toxicity of the brominated analogues within this class.
    Both DCA and TCA cause cardiac malformations in rats at high doses.

    1.3.4  Haloaldehydes and haloketones

         Chloral hydrate induces hepatic necrosis in rats at doses equal
    to or greater than 120 mg/kg of body weight per day. Its depressant
    effect on the central nervous system in humans is probably related to
    its metabolite trichloroethanol. Limited toxicity data are available
    for the other halogenated aldehydes and ketones. Chloroacetaldehyde
    exposure causes haematological effects in rats. Exposure of mice to
    1,1-dichloropropanone (1,1-DCPN), but not 1,3-dichloropropanone
    (1,3-DCPN), results in liver toxicity.

         Chloral hydrate was negative in most but not all bacterial tests
    for point mutations and in  in vivo studies on chromosomal damage.
    However, it has been shown that chloral hydrate may induce structural
    chromosomal aberrations  in vitro and  in vivo. Chloral hydrate has
    been reported to cause hepatic tumours in mice. It is not clear if it
    is the parent compound or its metabolites that are involved in the
    carcinogenic effect. The two chloral hydrate metabolites, TCA and DCA,
    have induced hepatic tumours in mice. 

         Some halogenated aldehydes and ketones are potent inducers of
    mutations in bacteria. Clastogenic effects have been reported for
    chlorinated propanones. Liver tumours were noted in a lifetime
    drinking-water study with chloroacetaldehyde. Other halogenated
    aldehydes, e.g., 2-chloropropenal, have been identified as tumour
    initiators in the skin of mice. The haloketones have not been tested
    for carcinogenicity in drinking-water. However, 1,3-DCPN acted as a
    tumour initiator in a skin carcinogenicity study in mice.

    1.3.5  Haloacetonitriles

         Testing of these compounds for toxicological effects has been
    limited to date. Some of the groups are mutagenic, but these effects
    do not relate well to the activity of the chemicals as tumour
    initiators in the skin. There are only very limited studies on the
    carcinogenicity of this class of substances. Early indications of
    developmental toxicity of members of this class appear to be largely
    attributable to the vehicle used in treatment.

    1.3.6  Halogenated hydroxyfuranone derivatives

         Based on experimental studies, the critical effects of MX appear
    to be its mutagenicity and carcinogenicity. Several  in vitro studies
    have revealed that MX is mutagenic in bacterial and mammalian test
    systems. MX caused chromosomal aberrations and induced DNA damage in
    isolated liver and testicular cells and sister chromatid exchanges in
    peripheral lymphocytes from rats exposed  in vivo. An overall

    evaluation of the mutagenicity data shows that MX is mutagenic
     in vitro and  in vivo. A carcinogenicity study in rats showed
    increased tumour frequencies in several organs.

    1.3.7  Chlorite

         The toxic action of chlorite is primarily in the form of
    oxidative damage to red blood cells at doses as low as 10 mg/kg of
    body weight. There are indications of mild neurobehavioural effects in
    rat pups at 5.6 mg/kg of body weight per day. There are conflicting
    data on the genotoxicity of chlorite. Chlorite does not increase
    tumours in laboratory animals in chronic exposure studies.

    1.3.8  Chlorate

         The toxicity of chlorate is similar to that of chlorite, but
    chlorate is less effective at inducing oxidative damage. It does not
    appear to be teratogenic or genotoxic  in vivo. There are no data
    from long-term carcinogenicity studies.

    1.3.9  Bromate

         Bromate causes renal tubular damage in rats at high doses. It
    induces tumours of the kidney, peritoneum and thyroid in rats at doses
    of 6 mg/kg of body weight and above in chronic studies. Hamsters are
    less sensitive, and mice are considerably less sensitive. Bromate is
    also genotoxic  in vivo in rats at high doses. Carcinogenicity
    appears to be secondary to oxidative stress in the cell.

    1.4  Epidemiological studies

    1.4.1  Cardiovascular disease

         Epidemiological studies have not identified an increased risk of
    cardiovascular disease associated with chlorinated or chloraminated
    drinking-water. Studies of other disinfectants have not been
    conducted.

    1.4.2  Cancer

         The epidemiological evidence is insufficient to support a causal
    relationship between bladder cancer and long-term exposure to
    chlorinated drinking-water, THMs, chloroform or other THM species. The
    epidemiological evidence is inconclusive and equivocal for an
    association between colon cancer and long-term exposure to chlorinated
    drinking-water, THMs, chloroform or other THM species. The information
    is insufficient to allow an evaluation of the observed risks for
    rectal cancer and risks for other cancers observed in single
    analytical studies.

         Various types of epidemiological studies have attempted to assess
    the cancer risks that may be associated with exposure to chlorinated
    drinking-water. Chloraminated drinking-water was considered in two
    studies. Several studies have attempted to estimate exposures to total

    THMs or chloroform and the other THM species, but the studies did not
    consider exposures to other DBPs or other water contaminants, which
    may differ for surface water and groundwater sources. One study
    considered the mutagenicity of drinking-water as measured by the
     Salmonella typhimurium assay. Assessments of possible cancer risks
    that may be associated with drinking-water disinfected with ozone or
    chlorine dioxide have not been performed.

         Ecological and death certificate-based case-control studies have
    provided hypotheses for further evaluation by analytical studies that
    consider an individual's exposure to drinking-water and possible
    confounding factors.

         Analytical studies have reported weak to moderate increased
    relative risks of bladder, colon, rectal, pancreatic, breast, brain or
    lung cancer associated with long-term exposure to chlorinated
    drinking-water. Single studies reported associations for pancreatic,
    breast or brain cancer; however, the evaluation of a possible causal
    relationship for epidemiological associations requires evidence from
    more than a single study. In one study, a small increased relative
    risk of lung cancer was associated with the use of surface water
    sources, but the magnitude of risk was too small to rule out residual
    confounding.

         A case-control study reported a moderately large association
    between rectal cancer and long-term exposure to chlorinated
    drinking-water or cumulative THM exposure, but cohort studies have
    found either no increased risk or a risk too weak to rule out residual
    confounding.

         Decreased bladder cancer risk was associated with increased
    duration of exposure to chloraminated drinking-water, but there is no
    biological basis for assuming a protective effect of chloraminated
    water.

         Although several studies found increased risks of bladder cancer
    associated with long-term exposure to chlorinated drinking-water and
    cumulative exposure to THMs, inconsistent results were reported among
    the studies for bladder cancer risks between smokers and non-smokers
    and between men and women. Estimated exposure to THMs was considered
    in three of these studies. In one study, no association was found
    between estimated cumulative exposure to THMs. In another study, a
    moderately strong increased relative risk was associated with
    increased cumulative exposure to THMs in men but not in women. The
    third study reported a weak increased relative risk associated with an
    estimated cumulative exposure of 1957-6425 g of THMs per litre-year;
    weak to moderate associations were also reported for exposure to THM
    concentrations greater than 24, greater than 49 and greater than 74
    g/litre. No increased relative risk of bladder cancer was associated
    with exposure to chlorinated municipal surface water supplies,
    chloroform or other THM species in a cohort of women, but the
    follow-up period of 8 years was very short, resulting in few cases for
    study.

         Because inadequate attention has been paid to assessing exposure
    to water contaminants in epidemiological studies, it is not possible
    to properly evaluate the increased relative risks that were reported.
    Specific risks may be due to other DBPs, mixtures of by-products or
    other water contaminants, or they may be due to other factors for
    which chlorinated drinking-water or THMs may serve as a surrogate.

    1.4.3  Adverse pregnancy outcomes

         Studies have considered exposures to chlorinated drinking-water,
    THMs or THM species and various adverse outcomes of pregnancy. A
    scientific panel recently convened by the US Environmental Protection
    Agency reviewed the epidemiological studies and concluded that the
    results of currently published studies do not provide convincing
    evidence that chlorinated water or THMs cause adverse pregnancy
    outcomes.

         Results of early studies are difficult to interpret because of
    methodological limitations or suspected bias.

         A recently completed but not yet published case-control study has
    reported moderate increased relative risks for neural tube defects in
    children whose mothers' residence in early pregnancy was in an area
    where THM levels were greater than 40 g/litre. Replication of the
    results in another area is required before this association can be
    properly evaluated. A previously conducted study in the same
    geographic area reported a similar association, but the study suffered
    from methodological limitations.

         A recently reported cohort study found an increased risk of early
    miscarriage associated with heavy consumption of water (five or more
    glasses of cold tapwater per day) containing high levels (>75
    g/litre) of THMs. When specific THMs were considered, only heavy
    consumption of water containing BDCM (>18 g/litre) was associated
    with a risk of miscarriage. As this is the first study to suggest an
    adverse reproductive effect associated with a brominated by-product, a
    scientific panel recommended that another study be conducted in a
    different geographic area to attempt to replicate these results and
    that additional efforts be made to evaluate exposures of the cohort to
    other water contaminants.

    1.5  Risk characterization

         It should be noted that the use of chemical disinfectants in
    water treatment usually results in the formation of chemical
    by-products, some of which are potentially hazardous. However, the
    risks to health from these by-products at the levels at which they
    occur in drinking-water are extremely small in comparison with the
    risks associated with inadequate disinfection. Thus, it is important
    that disinfection not be compromised in attempting to control such
    by-products.

    1.5.1  Characterization of hazard and dose-response

    1.5.1.1  Toxicological studies

    1)   Chlorine

         A WHO Working Group for the 1993  Guidelines for drinking-water 
     quality considered chlorine. This Working Group determined a
    tolerable daily intake (TDI) of 150 g/kg of body weight for free
    chlorine based on a no-observed-adverse-effect level (NOAEL) of
    approximately 15 mg/kg of body weight per day in 2-year studies in
    rats and mice and incorporating an uncertainty factor of 100 (10 each
    for intra- and interspecies variation). There are no new data that
    indicate that this TDI should be changed.

    2)   Monochloramine

         A WHO Working Group for the 1993  Guidelines for drinking-water 
     quality considered monochloramine. This Working Group determined a
    TDI of 94 g/kg of body weight based on a NOAEL of approximately 9.4
    mg/kg of body weight per day, the highest dose tested, in a 2-year
    bioassay in rats and incorporating an uncertainty factor of 100 (10
    each for intra- and interspecies variation). There are no new data
    that indicate that this TDI should be changed.

    3)   Chlorine dioxide

         The chemistry of chlorine dioxide in drinking-water is complex,
    but the major breakdown product is chlorite. In establishing a
    specific TDI for chlorine dioxide, data on both chlorine dioxide and
    chlorite can be considered, given the rapid hydrolysis to chlorite.
    Therefore, an oral TDI for chlorine dioxide is 30 g/kg of body
    weight, based on the NOAEL of 2.9 mg/kg of body weight per day for
    neurodevelopmental effects of chlorite in rats.

    4)   Trihalomethanes

         Cancer following chronic exposure is the primary hazard of
    concern for this class of DBPs. Because of the weight of evidence
    indicating that chloroform can induce cancer in animals only after
    chronic exposure to cytotoxic doses, it is clear that exposures to low
    concentrations of chloroform in drinking-water do not pose
    carcinogenic risks. The NOAEL for cytolethality and regenerative
    hyperplasia in mice was 10 mg/kg of body weight per day after
    administration of chloroform in corn oil for 3 weeks. Based on the
    mode of action evidence for chloroform carcinogenicity, a TDI of 10
    g/kg of body weight was derived using the NOAEL for cytotoxicity in
    mice and applying an uncertainty factor of 1000 (10 each for inter-
    and intraspecies variation and 10 for the short duration of the
    study). This approach is supported by a number of additional studies.
    This TDI is similar to the TDI derived in the 1998 WHO  Guidelines 
     for drinking-water quality, which was based on a 1979 study in which
    dogs were exposed for 7.5 years.

         Among the brominated THMs, BDCM is of particular interest because
    it produces tumours in rats and mice and at several sites (liver,
    kidneys, large intestine) after corn oil gavage. The induction of
    colon tumours in rats by BDCM (and by bromoform) is also interesting
    because of the epidemiological associations with colo-rectal cancer.
    BDCM and the other brominated THMs are also weak mutagens. It is
    generally assumed that mutagenic carcinogens will produce linear
    dose-response relationships at low doses, as mutagenesis is generally
    considered to be an irreversible and cumulative effect. 

         In a 2-year bioassay, BDCM given by corn oil gavage induced
    tumours (in conjunction with cytotoxicity and increased proliferation)
    in the kidneys of mice and rats at doses of 50 and 100 mg/kg of body
    weight per day, respectively. The tumours in the large intestine of
    the rat occurred after exposure to both 50 and 100 mg/kg of body
    weight per day. Using the incidence of kidney tumours in male mice
    from this study, quantitative risk estimates have been calculated,
    yielding a slope factor of 4.8  10-3 [mg/kg of body weight per
    day]-1 and a calculated dose of 2.1 g/kg of body weight per day for
    a risk level of 10-5. A slope factor of 4.2  10-3 [mg/kg of body
    weight per day]-1 (2.4 g/kg of body weight per day for a 10-5 risk)
    was derived based on the incidence of large intestine carcinomas in
    the male rat. The International Agency for Research on Cancer (IARC)
    has classified BDCM in Group 2B (possibly carcinogenic to humans).

         DBCM and bromoform were studied in long-term bioassays. In a
    2-year corn oil gavage study, DBCM induced hepatic tumours in female
    mice, but not in rats, at a dose of 100 mg/kg of body weight per day.
    In previous evaluations, it had been suggested that the corn oil
    vehicle may play a role in the induction of tumours in female mice. A
    small increase in tumours of the large intestine in rats was observed
    in the bromoform study at a dose of 200 mg/kg of body weight per day.
    The slope factors based on these tumours are 6.5  10-3 [mg/kg of
    body weight per day]-1 for DBCM, or 1.5 g/kg of body weight per day
    for a 10-5 risk, and 1.3  10-3 [mg/kg of body weight per day]-1 or
    7.7 g/kg of body weight per day for a 10-5 risk for bromoform.

         These two brominated THMs are weakly mutagenic in a number of
    assays, and they were by far the most mutagenic DBPs of the class in
    the GST-mediated assay system. Because they are the most lipophilic
    THMs, additional concerns about whether corn oil may have affected
    their bioavailability in the long-term studies should be considered. A
    NOAEL for DBCM of 30 mg/kg of body weight per day has been established
    based on the absence of histopathological effects in the liver of rats
    after 13 weeks of exposure by corn oil gavage. IARC has classified
    DBCM in Group 3 (not classifiable as to its carcinogenicity to
    humans). A TDI for DBCM of 30 g/kg of body weight was derived based
    on the NOAEL for liver toxicity of 30 mg/kg of body weight per day and
    an uncertainty factor of 1000 (10 each for inter- and intraspecies
    variation and 10 for the short duration of the study and possible
    carcinogenicity).

          Similarly, a NOAEL for bromoform of 25 mg/kg of body weight per
    day can be derived on the basis of the absence of liver lesions in
    rats after 13 weeks of dosing by corn oil gavage. A TDI for bromoform
    of 25 g/kg of body weight was derived based on this NOAEL for liver
    toxicity and an uncertainty factor of 1000 (10 each for inter- and
    intraspecies variation and 10 for the short duration of the study and
    possible carcinogenicity). IARC has classified bromoform in Group 3
    (not classifiable as to its carcinogenicity to humans).

    5)   Haloacetic acids

         The induction of mutations by DCA is very improbable at the low
    doses that would be encountered in chlorinated drinking-water. The
    available data indicate that DCA differentially affects the
    replication rates of normal hepatocytes and hepatocytes that have been
    initiated. The dose-response relationships are complex, with DCA
    initially stimulating division of normal hepatocytes. However, at the
    lower chronic doses used in animal studies (but still very high
    relative to those that would be derived from drinking-water), the
    replication rate of normal hepatocytes is eventually sharply
    inhibited. This indicates that normal hepatocytes eventually
    down-regulate those pathways that are sensitive to stimulation by DCA.
    However, the effects in altered cells, particularly those that express
    high amounts of a protein that is immunoreactive to a c-Jun antibody,
    do not seem to be able to down-regulate this response. Thus, the rates
    of replication in the pre-neoplastic lesions with this phenotype are
    very high at the doses that cause DCA tumours to develop with a very
    low latency. Preliminary data would suggest that this continued
    alteration in cell birth and death rates is also necessary for the
    tumours to progress to malignancy. This interpretation is supported by
    studies that employ initiation/promotion designs as well.

         On the basis of the above considerations, it is suggested that
    the currently available cancer risk estimates for DCA be modified by
    incorporation of newly developing information on its comparative
    metabolism and modes of action to formulate a biologically based
    dose-response model. These data are not available at this time, but
    they should become available within the next 2-3 years.

         The effects of DCA appear to be closely associated with doses
    that induce hepatomegaly and glycogen accumulation in mice. The
    lowest-observed-adverse-effect level (LOAEL) for these effects in
    an 8-week study in mice was 0.5 g/litre, corresponding to
    approximately 100 mg/kg of body weight per day, and the NOAEL was
    0.2 g/litre, or approximately 40 mg/kg of body weight per day. A TDI
    of 40 g/kg of body weight has been calculated by applying an
    uncertainty factor of 1000 to this NOAEL (10 each for inter- and
    intraspecies variation and 10 for the short duration of the study and
    possible carcinogenicity). IARC has classified DCA in Group 3 (not
    classifiable as to its carcinogenicity to humans).

         TCA is one of the weakest activators of the peroxisome
    proliferator activated receptor (PPAR) known. It appears to be only
    marginally active as a peroxisome proliferator, even in rats.

    Furthermore, treatment of rats with high levels of TCA in
    drinking-water does not induce liver tumours. These data strongly
    suggest that TCA presents little carcinogenic hazard to humans at the
    low concentrations found in drinking-water.

         From a broader toxicological perspective, the developmental
    effects of TCA are the end-point of concern. Animals appear to
    tolerate concentrations of TCA in drinking-water of 0.5 g/litre
    (approximately 50 mg/kg of body weight per day) with little or no
    signs of adverse effect. At 2 g/litre, the only sign of adverse effect
    appears to be hepatomegaly. Hepatomegaly is not observed in mice at
    doses of 0.35 g of TCA per litre in drinking-water, estimated to be
    equivalent to 40 mg/kg of body weight per day.

         In another study, soft tissue anomalies were observed at
    approximately 3 times the control rate at the lowest dose
    administered, 330 mg/kg of body weight per day. At this dose, the
    anomalies were mild and would clearly be in the range where
    hepatomegaly (and carcinogenic effects) would occur. Considering the
    fact that the PPAR interacts with cell signalling mechanisms that can
    affect normal developmental processes, a common mechanism underlying
    hepatomegaly and the carcinogenic effects and developmental effects of
    this compound should be considered.

         The TDI for TCA is based on a NOAEL estimated to be 40 mg/kg of
    body weight per day for hepatic toxicity in a long-term study in mice.
    Application of an uncertainty factor of 1000 (10 each for inter- and
    intraspecies variation and 10 for possible carcinogenicity) to the
    estimated NOAEL gives a TDI of 40 g/kg of body weight. IARC has
    classified TCA in Group 3 (not classifiable as to its carcinogenicity
    to humans).

         Data on the carcinogenicity of brominated acetic acids are too
    preliminary to be useful in risk characterization. Data available in
    abstract form suggest, however, that the doses required to induce
    hepatocarcinogenic responses in mice are not dissimilar to those of
    the chlorinated acetic acids. In addition to the mechanisms involved
    in the induction of cancer by DCA and TCA, it is possible that
    increased oxidative stress secondary to their metabolism might
    contribute to their effects.

         There are a significant number of data on the effects of
    dibromoacetic acid (DBA) on male reproduction. No effects were
    observed in rats at doses of 2 mg/kg of body weight per day for
    79 days, whereas an increased retention of step 19 spermatids was
    observed at 10 mg/kg of body weight per day. Higher doses led to
    progressively more severe effects, including marked atrophy of the
    seminiferous tubules with 250 mg/kg of body weight per day, which was
    not reversed 6 months after treatment was suspended. A TDI of 20 g/kg
    of body weight was determined by allocating an uncertainty factor of
    100 (10 each for inter- and intraspecies variation) to the NOAEL of
    2 mg/kg of body weight per day.

    6)   Chloral hydrate

         Chloral hydrate at 1 g/litre of drinking-water (166 mg/kg of body
    weight per day) induced liver tumours in mice exposed for 104 weeks.
    Lower doses have not been evaluated. Chloral hydrate has been shown to
    induce chromosomal anomalies in several  in vitro tests but has been
    largely negative when evaluated  in vivo. It is probable that the
    liver tumours induced by chloral hydrate involve its metabolism to TCA
    and/or DCA. As discussed above, these compounds are considered to act
    as tumour promoters. IARC has classified chloral hydrate in Group 3
    (not classifiable as to its carcinogenicity to humans).

         Chloral hydrate administered to rats for 90 days in
    drinking-water induced hepatocellular necrosis at concentrations of
    1200 mg/litre and above, with no effect being observed at 600 mg/litre
    (approximately 60 mg/kg of body weight per day). Hepatomegaly was
    observed in mice at doses of 144 mg/kg of body weight per day
    administered by gavage for 14 days. No effect was observed at 14.4
    mg/kg of body weight per day in the 14-day study, but mild
    hepatomegaly was observed when chloral hydrate was administered in
    drinking-water at 70 mg/litre (16 mg/kg of body weight per day) in a
    90-day follow-up study. The application of an uncertainty factor of
    1000 (10 each for inter- and intraspecies variation and 10 for the use
    of a LOAEL instead of a NOAEL) to this value gives a TDI of 16 g/kg
    of body weight.

    7)   Haloacetonitriles

         Without appropriate human data or an animal study that involves a
    substantial portion of an experimental animal's lifetime, there is no
    generally accepted basis for estimating carcinogenic risk from the
    HANs.

         Data developed in subchronic studies provide some indication of
    NOAELs for the general toxicity of dichloroacetonitrile (DCAN) and
    dibromoacetonitrile (DBAN). NOAELs of 8 and 23 mg/kg of body weight
    per day were identified in 90-day studies in rats for DCAN and DBAN,
    respectively, based on decreased body weights at the next higher doses
    of 33 and 45 mg/kg of body weight per day, respectively.

         A WHO Working Group for the 1993  Guidelines for drinking-water 
     quality considered DCAN and DBAN. This Working Group determined a
    TDI of 15 g/kg of body weight for DCAN based on a NOAEL of 15 mg/kg
    of body weight per day in a reproductive toxicity study in rats and
    incorporating an uncertainty factor of 1000 (10 each for intra- and
    interspecies variation and 10 for the severity of effects).
    Reproductive and developmental effects were observed with DBAN only at
    doses that exceeded those established for general toxicity (about 45
    mg/kg of body weight per day). A TDI of 23 g/kg of body weight was
    calculated for DBAN based on the NOAEL of 23 mg/kg of body weight per
    day in the 90-day study in rats and incorporating an uncertainty
    factor of 1000 (10 each for intra- and interspecies variation and 10
    for the short duration of the study). There are no new data indicating
    that these TDIs should be changed.

         LOAELs for trichloroacetonitrile (TCAN) of 7.5 mg/kg of body
    weight per day for embryotoxicity and 15 mg/kg of body weight per day
    for developmental effects were identified. However, later studies
    suggest that these responses were dependent upon the vehicle used. No
    TDI can be established for TCAN. 

         There are no data useful for risk characterization purposes for
    other members of the HANs.

    8)   MX

         The mutagen MX has recently been studied in a long-term study in
    rats in which some carcinogenic responses were observed. These data
    indicate that MX induces thyroid and bile duct tumours. An increased
    incidence of thyroid tumours was seen at the lowest dose of MX
    administered (0.4 mg/kg of body weight per day). The induction of
    thyroid tumours with high-dose chemicals has long been associated with
    halogenated compounds. The induction of thyroid follicular tumours
    could involve modifications in thyroid function or a mutagenic mode of
    action. A dose-related increase in the incidence of cholangiomas and
    cholangiocarcinomas was also observed, beginning at the low dose in
    female rats, with a more modest response in male rats. The increase in
    cholangiomas and cholangiocarcinomas in female rats was utilized to
    derive a slope factor for cancer. The 95% upper confidence limit for a
    10-5 lifetime risk based on the linearized multistage model was
    calculated to be 0.06 g/kg of body weight per day. 

    9)   Chlorite

         The primary and most consistent finding arising from exposure to
    chlorite is oxidative stress resulting in changes in the red blood
    cells. This end-point is seen in laboratory animals and, by analogy
    with chlorate, in humans exposed to high doses in poisoning incidents.
    There are sufficient data available with which to estimate a TDI for
    humans exposed to chlorite, including chronic toxicity studies and a
    two-generation reproductive toxicity study. Studies in human
    volunteers for up to 12 weeks did not identify any effect on blood
    parameters at the highest dose tested, 36 g/kg of body weight per
    day. Because these studies do not identify an effect level, they are
    not informative for establishing a margin of safety.

         In a two-generation study in rats, a NOAEL of 2.9 mg/kg of body
    weight per day was identified based on lower auditory startle
    amplitude, decreased absolute brain weight in the F1 and F2
    generations, and altered liver weights in two generations. Application
    of an uncertainty factor of 100 (10 each for inter- and intraspecies
    variation) to this NOAEL gives a TDI of 30 g/kg of body weight. This
    TDI is supported by the human volunteer studies.

    10)  Chlorate

         Like chlorite, the primary concern with chlorate is oxidative
    damage to red blood cells. Also like chlorite, 0.036 mg of chlorate
    per kg of body weight per day for 12 weeks did not result in any

    adverse effect in human volunteers. Although the database for chlorate
    is less extensive than that for chlorite, a recent well conducted
    90-day study in rats identified a NOAEL of 30 mg/kg of body weight per
    day based on thyroid gland colloid depletion at the next higher dose
    of 100 mg/kg of body weight per day. A TDI is not derived because a
    long-term study is in progress, which should provide more information
    on chronic exposure to chlorate.

    11)  Bromate

         Bromate is an active oxidant in biological systems and has been
    shown to cause an increase in renal tumours, peritoneal mesotheliomas
    and thyroid follicular cell tumours in rats and, to a lesser extent,
    hamsters, and only a small increase in kidney tumours in mice. The
    lowest dose at which an increased incidence of renal tumours was
    observed in rats was 6 mg/kg of body weight per day.

         Bromate has also been shown to give positive results for
    chromosomal aberrations in mammalian cells  in vitro and  in vivo 
    but not in bacterial assays for point mutation. An increasing body of
    evidence, supported by the genotoxicity data, suggests that bromate
    acts by generating oxygen radicals in the cell.

         In the 1993 WHO  Guidelines for drinking-water quality, the
    linearized multistage model was applied to the incidence of renal
    tumours in a 2-year carcinogenicity study in rats, although it was
    noted that if the mechanism of tumour induction is oxidative damage in
    the kidney, application of the low-dose cancer model may not be
    appropriate. The calculated upper 95% confidence interval for a 10-5
    risk was 0.1 g/kg of body weight per day.

         The no-effect level for the formation of renal cell tumours in
    rats is 1.3 mg/kg of body weight per day. If this is used as a point
    of departure from linearity and if an uncertainty factor of 1000 (10
    each for inter- and intraspecies variation and 10 for possible
    carcinogenicity) is applied, a TDI of 1 g/kg of body weight can be
    calculated. This compares with the value of 0.1 g/kg of body weight
    per day associated with an excess lifetime cancer risk of 10-5.

         At present, there are insufficient data to permit a decision on
    whether bromate-induced tumours are a result of cytotoxicity and
    reparative hyperplasia or a genotoxic effect.

         IARC has assigned potassium bromate to Group 2B (possibly
    carcinogenic to humans).

    1.5.1.2  Epidemiological studies

         Epidemiological studies must be carefully evaluated to ensure
    that observed associations are not due to bias and that the design is
    appropriate for an assessment of a possible causal relationship.
    Causality can be evaluated when there is sufficient evidence from
    several well designed and well conducted studies in different
    geographic areas. Supporting toxicological and pharmacological data

    are also important. It is especially difficult to interpret
    epidemiological data from ecological studies of disinfected
    drinking-water, and these results are used primarily to help develop
    hypotheses for further study.

         Results of analytical epidemiological studies are insufficient to
    support a causal relationship for any of the observed associations. It
    is especially difficult to interpret the results of currently
    published analytical studies because of incomplete information about
    exposures to specific water contaminants that might confound or modify
    the risk. Because inadequate attention has been paid to assessing
    exposures to water contaminants in epidemiological studies, it is not
    possible to properly evaluate the increased relative risks that were
    reported. Risks may be due to other water contaminants or to other
    factors for which chlorinated drinking-water or THMs may serve as a
    surrogate.

    1.5.2  Characterization of exposure

    1.5.2.1  Occurrence of disinfectants and disinfectant by-products

         Disinfectant doses of several milligrams per litre are typically
    employed, corresponding to doses necessary to inactivate
    microorganisms (primary disinfection) or doses necessary to maintain a
    residual in the distribution system (secondary disinfection).

         A necessary ingredient for an exposure assessment is DBP
    occurrence data. Unfortunately, there are few published international
    studies that go beyond case-study or regional data.

         Occurrence data suggest, on average, exposure to about 35-50 g
    of total THMs per litre in chlorinated drinking-water, with chloroform
    and BDCM being the first and second most dominant species. Exposure to
    total HAAs can be approximated by a total HAA concentration (sum of
    five species) corresponding to about one-half of the total THM
    concentration (although this ratio can vary significantly); DCA and
    TCA are the first and second most dominant species. In waters with a
    high bromide to TOC ratio or a high bromide to chlorine ratio, greater
    formation of brominated THMs and HAAs can be expected. When a
    hypochlorite solution (versus chlorine gas) is used, chlorate may also
    occur during chlorination.

         DBP exposure in chloraminated water is a function of the mode of
    chloramination, with the sequence of chlorine followed by ammonia
    leading to the formation of (lower levels of) chlorine DBPs (i.e.,
    THMs and HAAs) during the free-chlorine period; however, the
    suppression of chloroform and TCA formation is not paralleled by a
    proportional reduction in DCA formation.

         All factors being equal, bromide concentration and ozone dose are
    the best predictors of bromate formation during ozonation, with about
    a 50% conversion of bromide to bromate. A study of different European
    water utilities showed bromate levels in water leaving operating water

    treatment plants ranging from less than the detection limit (2
    g/litre) up to 16 g/L. The brominated organic DBPs formed upon
    ozonation generally occur at low levels. The formation of chlorite can
    be estimated by a simple percentage (50-70%) of the applied chlorine
    dioxide dose.

    1.5.2.2  Uncertainties of water quality data

         A toxicological study attempts to extrapolate a laboratory
    (controlled) animal response to a potential human response; one
    possible outcome is the estimation of cancer risk factors. An
    epidemiological study attempts to link human health effects (e.g.,
    cancer) to a causative agent or agents (e.g., a DBP) and requires an
    exposure assessment. 

         The chemical risks associated with disinfected drinking-water are
    potentially based on several routes of exposure: (i) ingestion of DBPs
    in drinking-water; (ii) ingestion of chemical disinfectants in
    drinking-water and the concomitant formation of DBPs in the stomach;
    and (iii) inhalation of volatile DBPs during showering. Although the
     in vivo formation of DBPs and the inhalation of volatile DBPs may be
    of potential health concern, the following discussion is based on the
    premise that the ingestion of DBPs present in drinking-water is the
    most significant route of exposure.

         Human exposure is a function of both DBP concentration and
    exposure time. More specifically, human health effects are a function
    of exposure to complex mixtures of DBPs (e.g., THMs versus HAAs,
    chlorinated versus brominated species) that can change
    seasonally/temporally (e.g., as a function of temperature, nature and
    concentration of NOM) and spatially (i.e., throughout a distribution
    system). Each individual chemical disinfectant can form a mixture of
    DBPs; combinations of chemical disinfectants can form even more
    complex mixtures. Upon their formation, most DBPs are stable, but some
    may undergo transformation by, for example, hydrolysis. In the absence
    of DBP data, surrogates such as chlorine dose (or chlorine demand),
    TOC (or ultraviolet absorbance at 254 nm [UVA254]) or bromide can be
    used to indirectly estimate exposure. While TOC serves as a good
    surrogate for organic DBP precursors, UVA254 provides additional
    insight into NOM characteristics, which can vary geographically. Two
    key water quality variables, pH and bromide, have been identified as
    significantly affecting the type and concentrations of DBPs that are
    produced.

         An exposure assessment should first attempt to define the
    individual types of DBPs and resultant mixtures likely to form, as
    well as their time-dependent concentrations, as affected by their
    stability and transport through a distribution system. For
    epidemiological studies, some historical databases exist for
    disinfectant (e.g., chlorine) doses, possibly DBP precursor (e.g.,
    TOC) concentrations and possibly total THM (and, in some cases, THM
    species) concentrations. In contrast to THMs, which have been
    monitored over longer time frames because of regulatory scrutiny,
    monitoring data for HAAs (and HAA species), bromate and chlorite are

    much more recent and hence sparse. However, DBP models can be used to
    simulate missing or past data. Another important consideration is
    documentation of past changes in water treatment practice. 

    1.5.2.3  Uncertainties of epidemiological data

         Even in well designed and well conducted analytical studies,
    relatively poor exposure assessments were conducted. In most studies,
    duration of exposure to disinfected drinking-water and the water
    source were considered. These exposures were estimated from
    residential histories and water utility or government records. In only
    a few studies was an attempt made to estimate a study participant's
    water consumption and exposure to either total THMs or individual
    species of THMs. In only one study was an attempt made to estimate
    exposures to other DBPs. In evaluating some potential risks, i.e.,
    adverse outcomes of pregnancy, that may be associated with relatively
    short term exposures to volatile by-products, it may be important to
    consider the inhalation as well as the ingestion route of exposure
    from drinking-water. In some studies, an effort was made to estimate
    both by-product levels in drinking-water for etiologically relevant
    time periods and cumulative exposures. Appropriate models and
    sensitivity analysis such as Monte Carlo simulation can be used to
    help estimate these exposures for relevant periods.

         A major uncertainty surrounds the interpretation of the observed
    associations, as exposures to a relatively few water contaminants have
    been considered. With the current data, it is difficult to evaluate
    how unmeasured DBPs or other water contaminants may have affected the
    observed relative risk estimates. 

         More studies have considered bladder cancer than any other
    cancer. The authors of the most recently reported results for bladder
    cancer risks caution against a simple interpretation of the observed
    associations. The epidemiological evidence for an increased relative
    risk of bladder cancer is not consistent -- different risks are
    reported for smokers and non-smokers, for men and women, and for high
    and low water consumption. Risks may differ among various geographic
    areas because the DBP mix may be different or because other water
    contaminants are also present. More comprehensive water quality data
    must be collected or simulated to improve exposure assessments for
    epidemiological studies.
	
	

    2.  CHEMISTRY OF DISINFECTANTS AND DISINFECTANT BY-PRODUCTS

    2.1  Background

         The use of chlorine (Cl2) as a water disinfectant has come under
    scrutiny because of its potential to react with natural organic matter
    (NOM) and form chlorinated disinfectant by-products (DBPs). Within
    this context, NOM serves as the organic DBP precursor, whereas bromide
    ion (Br-) serves as the inorganic precursor. Treatment strategies
    generally available to water systems exceeding drinking-water
    standards include removing DBP precursors and using alternative
    disinfectants for primary and/or secondary (distribution system)
    disinfection. Alternative disinfectant options that show promise are
    chloramines (NH2Cl, monochloramine), chlorine dioxide (ClO2) and
    ozone (O3). While ozone can serve as a primary disinfectant only and
    chloramines as a secondary disinfectant only, both chlorine and
    chlorine dioxide can serve as either primary or secondary
    disinfectants.

         Chloramine presents the significant advantage of virtually
    eliminating the formation of chlorination by-products and, unlike
    chlorine, does not react with phenols to create taste- and
    odour-causing compounds. However, the required contact time for
    inactivation of viruses and  Giardia cysts is rarely obtainable by
    chloramine post-disinfection at existing water treatment facilities
    (monochloramine is significantly less biocidal than free chlorine).
    More recently, the presence of nitrifying bacteria and nitrite
    (NO2-) and nitrate (NO3-) production in chloraminated distribution
    systems as well as the formation of organic chloramines have raised
    concern. 

         The use of chlorine dioxide, like chloramine, can reduce the
    formation of chlorinated by-products during primary disinfection.
    However, production of chlorine dioxide, its decomposition and
    reaction with NOM lead to the formation of by-products such as
    chlorite (ClO2-), a compound that is of health concern.

         If used as a primary disinfectant followed by a chloramine
    residual in the distribution system, ozone can eliminate the need for
    contact between DBP precursors and chlorine. Ozone is known to react
    both with NOM to produce organic DBPs such as aldehydes and increase
    levels of assimilable organic carbon and with bromide ion to form
    bromate.

         A thorough understanding of the mechanisms of DBP formation
    allows microbial inactivation goals and DBP control goals to be
    successfully balanced. This chapter examines a range of issues
    affecting DBP formation and control to provide guidance to utilities
    considering the use of various disinfecting chemicals to achieve
    microbial inactivation with DBP control.

    2.2  Physical and chemical properties of common disinfectants and
         inorganic disinfectant by-products

         The important physical and chemical properties of commonly used
    disinfectants and inorganic DBPs are summarized in Table 1.

    2.2.1  Chlorine

         Chlorine, a gas under normal pressure and temperature, can be
    compressed to a liquid and stored in cylindrical containers. Because
    chlorine gas is poisonous, it is dissolved in water under vacuum, and
    this concentrated solution is applied to the water being treated. For
    small plants, cylinders of about 70 kg are used; for medium to large
    plants, tonne containers are common; and for very large plants,
    chlorine is delivered by railway tank cars or road (truck) tankers.
    Chlorine is also available in granular or powdered form as calcium
    hypochlorite (Ca(OCl)2) or in liquid form as sodium hypochlorite
    (NaOCl; bleach).

         Chlorine is used in the form of gaseous chlorine or hypochlorite
    (OCl-). In either form, it acts as a potent oxidizing agent and often
    dissipates in side reactions so rapidly that little disinfection is
    accomplished until amounts in excess of the chlorine demand have been
    added. As an oxidizing agent, chlorine reacts with a wide variety of
    compounds, in particular those that are considered reducing agents
    (hydrogen sulfide [H2S], manganese(II), iron(II), sulfite [SO32-],
    Br-, iodide [I-], nitrite). From the point of view of DBP formation
    and disinfection, these reactions may be important because they may be
    fast and result in the consumption of chlorine.

         Chlorine gas hydrolyses in water almost completely to form
    hypochlorous acid (HOCl):

              Cl2 + H2O -> HOCl + H+ + Cl-

         The hypochlorous acid dissociates into hydrogen ions (H+) and
    hypochlorite ions in the reversible reaction:

              HOCl <-> H+ + OCl-

         Hypochlorous acid is a weak acid with a p Ka of approximately
    7.5 at 25C. Hypochlorous acid, the prime disinfecting agent, is
    therefore dominant at a pH below 7.5 and is a more effective
    disinfectant than hypochlorite ion, which dominates above pH 7.5.

         The rates of the decomposition reactions of chlorine increase as
    the solution becomes more alkaline, and these reactions can
    theoretically produce chlorite and chlorate (ClO3-); they occur
    during the electrolysis of chloride (Cl-) solutions when the anodic
    and cathodic compartments are not separated, in which case the
    chlorine formed at the anode can react with the alkali formed at the
    cathode. On the other hand, hypochlorous acid/hypochlorite (or
    hypobromous acid/hypobromite, HOBr/OBr-) can be formed by the action
    of chlorine (or bromine) in neutral or alkaline solutions. The


        Table 1. Physical and chemical properties of commonly used disinfectants and inorganic disinfectant by-products

                                                                                                                              

    Chemicala       Eo (V)b      Oxidation number    lambamax (nm)c      e (mol-1 litre-1 cm-1)d     p epsilono e    pKa f
                                 of Cl or Br
                                                                                                                              

    HOCl/Cl-        +1.49        +1                  254                 60                          +25.2           7.5
                                                     292 (OCl-)          419
    ClO2/ClO2-      +0.95        +4                  359                 1250                        +16.1           -
    NH2Cl           -            +1                  245                 416                         -               -
    O3/O2           +2.07        -                   254                 3200                        +35.0           
    HOBr/Br-        +1.33        +1                  330                 50                          +22.5           8.7
    ClO2-/Cl-       +0.76        +3                  262                 -                           +12.8           1.96
    ClO3-/Cl-       +0.62        +5                  360                 -                           +10.5           1.45
    BrO3-/Br-       +0.61        +5                  195                 -                                           0.72
                                                                                                                              

    a  Half-cell reactants/products.
    b  Eo = standard electrode potential (redox potential) in water at 25 C. The oxidation-reduction state of an aqueous 
       environment at equilibrium can be stated in terms of its redox potential. In the chemistry literature, this is generally 
       expressed in volts, E, or as the negative logarithm of the electron activity, p epsilon. When p epsilon is large, the 
       electron activity is low and the system tends to be an oxidizing one: i.e., half-reactions tend to be driven to the left. 
       When p epsilon is small, the system is reducing, and reactions tend to be driven to the right.
    c  lambdamax = maximum absorbance wavelength of that particular solution in nm.
    d  e = molar absorptivity (molar extinction coefficient), in mol-1 litre-1 cm-1. This can be used for quantitative 
       determination of the various species of chemicals and is the only direct physical measurement. There is often some 
       background absorbance that may interfere with the measurement in natural waters that should be considered.
    e  p epsilono = - log {e-}  where {e-} = electron activity.
    f  pKa = negative logarithm of the acid ionization constant (e.g., at pH 7.5, the molar concentration of HOCl is same as that 
       of OCl-). As this parameter is dependent upon temperature, the values listed were determined at 25 C.
    

    decomposition of hypohalites (XO-) is favoured in alkaline solutions
    (2XO- -> X- + XO2-) and is such that there is no longer any
    domain of thermodynamic stability for the hypohalite ions. These
    oxyhalites are further converted to stable oxyhalates as follows:

              XO- + XO2- -> X- + XO3-

         Another reaction that occurs in waters containing bromide ion and
    hypochlorite is the production of hypobromous acid:

              HOCl + Br- -> HOBr + Cl-

    This reaction is irreversible, and the product hypobromous acid is a
    better halogenating agent than hypochlorous acid and interferes with
    common analytical procedures for free chlorine. The presence of
    bromide in hypochlorite solutions can ultimately lead to the formation
    of bromate (BrO3-).

         Hypobromous acid is a weak acid (p Ka = 8.7); like
    hypochlorite, hypobromite is metastable. In alkaline solution, it
    decomposes to give bromate and bromide:

              3OBr- -> BrO3- + 2Br-

         Bromic acid (HBrO3) is a strong acid (p Ka = 0.7). Bromic acid
    and bromate can be obtained by the electrolytic oxidation of bromide
    solutions or bromine water using chlorine. Bromic acid and bromate are
    powerful oxidizing agents, but the speed of their oxidation reactions
    is generally slow (Mel et al., 1953).

    2.2.2  Chlorine dioxide

         Chlorine dioxide is one of the few compounds that exists almost
    entirely as monomeric free radicals. Concentrated chlorine dioxide
    vapour is potentially explosive, and attempts to compress and store
    this gas, either alone or in combination with other gases, have been
    commercially unsuccessful. Because of this, chlorine dioxide, like
    ozone, must be manufactured at the point of use. Chlorine dioxide in
    water does not hydrolyse to any appreciable extent. Neutral or acidic
    dilute aqueous solutions are quite stable if kept cool, well sealed
    and protected from sunlight.

         Chlorine dioxide represents an oxidation state (+4) intermediate
    between those of chlorite (+3) and chlorate (+5). No acid or ion of
    the same degree of oxidation is known. Chlorine dioxide is a powerful
    oxidizing agent that can decompose to chlorite; in the absence of
    oxidizable substances and in the presence of alkali, it dissolves in
    water, decomposing with the slow formation of chlorite and chlorate:

              2ClO2 + H2O -> ClO2- + ClO3- + 2H+

         Chlorine dioxide has an absorption spectrum with a maximum at 359
    nm, with a molar absorptivity of 1250 mol-1 litre-1 cm-1. This
    extinction coefficient is independent of temperature, pH, chloride and

    ionic strength. Chlorine dioxide is readily soluble in water, forming
    a greenish-yellow solution. It can be involved in a variety of redox
    reactions, such as oxidation of iodide ion, sulfide ion, iron(II) and
    manganese(II). When chlorine dioxide reacts with aqueous contaminants,
    it is usually reduced to chlorite ion. The corresponding electron
    transfer reactions are comparable to those occurring when singlet
    oxygen acts as an oxidant (Tratnyek & Hoigne, 1994).

         Bromide (in the absence of sunlight) is not oxidized by chlorine
    dioxide. Therefore, water treatment with chlorine dioxide will not
    transform bromide ion into hypobromite and will not give rise to the
    formation of bromoform (CHBr3) or bromate. This is an important
    difference between the use of chlorine dioxide as an oxidant and the
    use of chlorine or ozone as an oxidant.

    2.2.3  Ozone

         Ozone is a strong oxidizing agent ( Eo = 2.07 V). Oxidation
    reactions initiated by ozone in water are generally rather complex; in
    water, only part of the ozone reacts directly with dissolved solutes.
    Another part may decompose before reaction. Such decomposition is
    catalysed by hydroxide ions (OH-) and other solutes. Highly reactive
    secondary oxidants, such as hydroxyl radicals (OH.), are thereby
    formed. These radicals and their reaction products can additionally
    accelerate the decomposition of ozone. Consequently, radical-type
    chain reactions may occur, which consume ozone concurrently with the
    direct reaction of ozone with dissolved organic material.

         Many oxidative applications of ozone have been developed,
    including disinfection, control of algae, removal of tastes and
    odours, removal of colour, removal of iron and manganese,
    microflocculation, removal of turbidity by oxidative flocculation,
    removal of organics by oxidation of phenols, detergents and some
    pesticides, partial oxidation of dissolved organics and control of
    halogenated organic compounds. For disinfection and for oxidation of
    many organic and inorganic contaminants in drinking-water, the
    kinetics of ozone reactions are favourable; on the other hand, for
    many difficult-to-oxidize organic compounds, such as chloroform
    (CHCl3), the kinetics of ozone oxidation are very slow (Hoigne et
    al., 1985).

    2.2.4  Chloramines

         Monochloramine has much higher CT values1 than free chlorine
    and is therefore a poor primary disinfectant. Additionally, it is a
    poor oxidant and is not effective for taste and odour control or for
    oxidation of iron and manganese. However, because of its persistence,


              

    1 The CT value is the product of the disinfectant concentration  C 
    in mg/litre and the contact time  T in minutes required to inactivate
    a specified percentage (e.g., 99%) of microorganisms.

    it is an attractive secondary disinfectant for the maintenance of a
    stable distribution system residual. The use of disinfectants such as
    ozone or chlorine dioxide combined with chloramines as a secondary
    disinfectant appears to be attractive for minimizing DBP formation
    (Singer, 1994b).

         Monochloramine is the only useful ammonia-chloramine
    disinfectant. Dichloramine (NHCl2) and nitrogen trichloride (NCl3)
    are too unstable to be useful and highly malodorous. Conditions
    practically employed for chloramination are designed to produce only
    monochloramine.

    2.3  Analytical methods for disinfectant by-products and disinfectants

         Analytical methods for various DBPs and their detection limits
    are summarized in Table 2. Methods for disinfectants are summarized in
    APHA (1995).

    2.3.1  Trihalomethanes, haloacetonitriles, chloral hydrate,
           chloropicrin and haloacetic acids

         Gas chromatographic (GC) techniques are generally employed
    for organic DBPs. Detection and quantification of haloacetonitriles
    (HANs) and chloral hydrate in chlorinated natural waters are
    complicated by (i) hydrolysis of dihaloacetonitriles and chloral
    hydrate to dihaloacetic acids and chloroform, respectively; (ii)
    degradation of HANs by dechlorinating agents such as sodium sulfite
    and sodium thiosulfate; (iii) low purge efficiency for the HANs and
    chloral hydrate in the purge-and-trap technique; and (iv) low
    extraction efficiency for chloral hydrate with pentane in the
    liquid-liquid extraction normally used. Although chloral hydrate is
    not efficiently extracted from water with pentane, it can be extracted
    with an efficiency of approximately 36% when the ratio by volume of
    methyl  tert-butyl ether (MTBE) to water is 1 : 5 (Amy et al., 1998).
    MTBE quantitatively extracts HANs, trihalomethanes (THMs), chloral
    hydrate and chloropicrin, permitting simultaneous analysis for all of
    these DBPs. Chloral hydrate decomposes on packed columns to
    trichloroacetaldehyde, resulting in considerable band broadening,
    although this does not appear to be a significant problem with DB-1
    and DB-5 columns.

         The extraction of THMs can be accomplished using MTBE (EPA Method
    551) or pentane. Method 551 also permits simultaneous extraction and
    measurement of chloral hydrate, HANs, THMs, chloropicrin and
    haloketones. The pentane method can be used to extract THMs, HANs,
    haloketones and chloropicrin but not chloral hydrate in the same run
    (APHA, 1995).

         The haloacetic acid (HAA) analytical method involves using an
    acidic salted ether or acidic methanol liquid-liquid extraction,
    requiring esterification with diazomethane prior to analysis on a gas
    chromatograph equipped with an electron capture detector (ECD). THMs
    and HANs can be analysed by extraction with pentane prior to analysis
    on a capillary-column GC equipped with an ECD. The analysis of


        Table 2. Summary of analytical methods for various DBPs and their minimum detection limits

                                                                                                                                        

    DBPs                Analytical method                    APHAa       Minimum detection    Major            References
                                                             method      limit (g/litre)     interferences
                                                                                                                                        

    THMs                MTBE extraction                      -           0.4                  None             AWWARF (1991)
                        Pentane extraction                   6232B       0.1
    HAAs                Salted MTBE extraction and           6233B       0.5-1.0              None             AWWARF (1991)
                        derivatization with diazomethane
    HANs                Pentane extraction                   6232B       0.05                 None             Koch et al. (1988)
    Cyanogen chloride   MTBE extraction                      6233A       0.5                  None             AWWARF (1991)
    Chloramine          Derivatization with                  -           -                    None             Lukasewycz et al. (1989)
                        2-mercaptobenzothiazole
    Haloketonesb        Pentane extraction                   6232B       0.2                  None             Krasner et al. (1995)
                                                                                                               AWWARF (1991)
    Chloral hydrate     MTBE extraction                      -           0.5                  None             AWWARF (1991)
    Aldehydes           Extraction with hexane and           -           1.0                  PFBHA sulfate    Sclimenti et al. (1990)
                        derivatization with PFBHA
    Bromate             Ion chromatography (H3BO3/NaOH)      4500        2.0c                 Cl-              Siddiqui et al. (1996a)
                                                                                                               Krasner et al. (1993)
                                                                         0.2                                   Weinberg et al. (1998)
    Chlorate            Ion chromatography (H3BO3/NaOH)      4500        5                    Cl-, acetate     Siddiqui (1996)
    Chlorite            Ion chromatography (NaHCO3/Na2CO3)   4500        10                   Cl-, acetate     AWWARF (1991)
    TOC                 UV/persulfate or combustion          5310        200                  Metals           APHA (1995)
                                                                                                                                        

    a  American Public Health Association.
    b  Sum of 1,1-DCPN and 1,1,1-TCPN.
    c  1.0 g/litre with high-capacity column.
    

    cyanogen compounds involves extraction with MTBE prior to injection
    into GC-ECD. Aldehydes require derivatization with
     O-(2,3,4,5,6-pentafluorobenzyl)-hydroxylamine (PFBHA) (to form an
    oxime), extraction with hexane and GC-ECD analysis [(C6F5)-CH2ONH2
    + RCHO -> (C6F5)-CH2ON=CHR + H2O]. It should be noted that PFBHA
    peaks are very large relative to other peaks in the chromatogram from
    a purge-and-trap system, whereas the peaks are comparable to other
    peaks in a GC-ECD chromatogram (Trehy et al., 1986).

    2.3.2  Inorganic disinfectant by-products

         An ion chromatography (IC) method (EPA Method 300) has been
    developed to determine inorganic by-products. The elution order is
    fluoride, chlorite, bromate, chloride, nitrite, bromide, chlorate,
    nitrate and sulfate ion. The eluent is a carbonate buffer.
    Ethylenediamine is used to preserve chlorite samples and to minimize
    the potential for chlorite ion reaction on the IC separating column.
    EPA Method 300 involves measurement by an IC system using a separating
    column (e.g., Ion Pac AS9-SC) fitted with an anion micromembrane
    suppressor column. An eluent containing 2.0 mmol of sodium carbonate
    (Na2CO3) per litre / 0.75 mmol of sodium bicarbonate (NaHCO3) per
    litre is used for bromide determination, and an eluent containing 40
    mmol of boric acid (H3BO3) per litre / 20 mmol of sodium hydroxide
    (NaOH) per litre is used for bromate and chlorate determination. The
    analytical minimum detection limits for bromate and chlorate using a
    borate eluent have been reported as 2 g/litre and 5 g/litre,
    respectively (Siddiqui, 1996; Siddiqui et al., 1996a). For samples
    with high chloride ion content, a silver cartridge can be used to
    remove chloride prior to IC analysis to minimize its interference with
    bromate measurement. It should be noted that for natural sources and
    waters with high total organic carbon (TOC) levels, detection limits
    will be slightly different because of the masking effect of NOM and
    high concentrations of carbonate/bicarbonate ions that may interfere
    with bromate/chlorate measurement.

    2.3.3  Total organic carbon and UV absorbance at 254 nm

         TOC is the primary surrogate parameter for the measurement of NOM
    in water supplies. Several investigators have reported that the
    ultraviolet (UV)/persulfate oxidation method underestimates the TOC
    concentration in natural waters as compared with the combustion method
    because of the inability of the persulfate method to oxidize highly
    polymerized organic matter. It is generally assumed that the
    calibration of a TOC analyser with a potassium hydrogen phthalate
    standard is sufficient for the measurement of TOC in natural waters,
    but potassium hydrogen phthalate has a simple molecular structure and
    is easy to oxidize. Dissolved organic carbon (DOC) is operationally
    defined by a (0.45-m) filtration step. UV absorbance at 254 nm
    (UVA254) is used to describe the type and character of NOM, whereas
    TOC describes just the amount of NOM.

    2.3.4  Chloramines

         Knowledge of the amine content of the water during water
    treatment processes involving chloramination is important to define
    more adequately the content of a matrix described only as a combined
    chlorine residual. The presence of organic nitrogen and the
    instability of many organic chloramines continue to challenge the
    analyst. Lukasewycz et al. (1989) developed a technique for the
    analysis of chloramines and organic chloramines present in water using
    2-mercaptobenzothiazole as a derivatizing agent. The resulting
    sulfanilamides are stable and can be conveniently analysed by
    high-performance liquid chromatography (HPLC) using UV or
    electrochemical detection. This method appears to be superior to the
    use of diazotization or phenylarsine oxide as a method of detection.
    Organic chloramines are much weaker disinfectants than inorganic
    monochloramine but are indistinguishable by the common analytical
    methods.

    2.4  Mechanisms involved in the formation of disinfectant by-products

    2.4.1  Chlorine reactions

         Chlorine reacts with humic substances (dissolved organic matter)
    present in most water supplies, forming a variety of halogenated DBPs,
    such as THMs, HAAs, HANs, chloral hydrate and chloropicrin, as
    follows:

              HOCl + DOC -> DBPs

         It is generally accepted that the reaction between chlorine and
    humic substances, a major component of NOM, is responsible for the
    production of organochlorine compounds during drinking-water
    treatment. Humic and fulvic acids show a high reactivity towards
    chlorine and constitute 50-90% of the total DOC in river and lake
    waters (Thurman, 1985). Other fractions of the DOC comprise the
    hydrophilic acids (up to 30%), carbohydrates (10%), simple carboxylic
    acids (5%) and proteins/amino acids (5%). The reactivity of
    carbohydrates and carboxylic acids towards chlorine is low, and they
    are not expected to contribute to the production of organochlorine
    compounds. However, hydrophilic acids such as citric acid and amino
    acids will react with chlorine to produce chloroform and other
    products and may contribute towards total organochlorine production
    (Larson & Rockwell, 1979).

         Free chlorine reacts with water constituents by three general
    pathways: oxidation, addition and substitution (Johnson & Jensen,
    1986). Chlorine can undergo an addition reaction if the organic
    compound has a double bond. For many compounds with double bonds, this
    reaction is too slow to be of importance in water treatment. The
    oxidation reactions with water constituents such as carbohydrates or
    fatty acids (e.g., oleic acid) are generally slow.

         Most chlorine DBPs are formed through oxidation and substitution
    reactions. THMs have the general formula CHX3, where X can be Cl or
    Br. Chloroform may be produced through a series of reactions with
    functional groups of humic substances. The major functional groups of
    humic substances include acetyl, carboxyl, phenol, alcohol, carbonyl
    and methoxyl. The reactions proceed much more rapidly at high pH than
    at low pH.

         Rook (1977) proposed resorcinol structures to be the major
    precursor structure in humic material for chloroform formation. In
    accordance with this hypothesis in the chlorination of terrestrial and
    aquatic humic substances, a series of intermediates were detected that
    contained a trichloromethyl group and that could be converted to
    chloroform by further oxidation or substitution reactions (Stevens et
    al., 1976).

         However, the production of chlorinated compounds such as
    dichloropropanedoic acid, 2,2-dichlorobutanedoic acid, cyanogen
    chloride (CNCl), HANs or the cyano-substituted acids cannot be
    explained on the basis of resorcinol structures, and possible
    production pathways require protein-type precursors (De Leer et al.,
    1986). The reaction pathway for amino acids involves initial rapid
    formation of the monochloramine and dichloramine, which can react
    further to form aldehyde or HANs, respectively. Trehy et al. (1986)
    demonstrated the formation of chloral hydrate along with HANs after
    chlorination of amino acids by substitution reactions, and aldehydes
    were shown to be the oxidation products. Luknitskii (1975) provided a
    detailed chemistry of chloral hydrate formation.

         Christman et al. (1983) also identified chloroform, chloral
    hydrate, dichloroacetic acid (DCA), trichloroacetic acid (TCA) and
    2,2-dichlorobutanedoic acid as the major products, accounting for 53%
    of the total organic halogen (TOX). A number of other minor products
    have been detected, including several chlorinated alkanoic acids and
    non-chlorinated benzene carboxylic acids. De Leer et al. (1985)
    extended these studies to incorporate chloroform intermediates,
    chlorinated aromatic acids and cyano-compounds as potential products
    in drinking-water. The presence of unhalogenated aldehydes and HANs in
    chlorinated natural waters can be attributed in part to the presence
    of amino acids or peptides in natural waters. Humic acids may also
    contribute to the presence of amino acids in natural waters, as they
    have amino acids associated with them either in a free or in a
    combined form. Several studies regarding the chlorination of amino
    acids have shown that the primary amino group on the amino acids can
    be converted to either an aldehyde or a nitrile group (Morris et al.,
    1980; Isaac & Morris, 1983). These studies indicate that with an
    equimolar amount of halogenating agent, the major product is an
    aldehyde. However, if an excess of halogenating agent is added, then
    the corresponding nitrile can also be formed, with the ratio of the
    aldehyde to nitrile formed increasing with pH.

         Many treated waters contain not only chlorinated but also
    brominated compounds, such as bromoform. These compounds form because
    aqueous chlorine converts bromide in the water to hypobromous acid.

    The bromine can then react with the organic matter in the same way as
    hypochlorous acid to form various bromochlorinated DBPs. However,
    compared with hypochlorous acid, hypobromous acid is a weaker oxidant
    and stronger halogenating agent.

         Chlorate, an inorganic by-product of chlorine, is formed in
    concentrated hypochlorite solutions during their production and
    storage through the following reactions (Gordon et al., 1997):

              OCl- + OCl- -> ClO2- + Cl-

              OCl- + ClO2- -> ClO3- + Cl-

         The first reaction proceeds at a much slower rate and is rate
    limiting, hence the generally observed second-order kinetics. Sodium
    hypochlorite is stored at pH greater than 12 to prevent rapid
    decomposition, and most of the sodium hypochlorite is present as
    hypochlorite ion. The average rate constant for the formation of
    chlorate is 85  10-5 mol-1 litre-1 d-1 (Gordon et al., 1995).

    2.4.2  Chlorine dioxide reactions

         The major chlorine dioxide by-products of concern are chlorite
    and chlorate. Chlorine dioxide reacts generally as an electron
    acceptor, and hydrogen atoms present in activated organic C-H or N-H
    structures are thereby not substituted by chlorine (Hoigne & Bader,
    1994). Moreover, in contrast to chlorine, chlorine dioxide's
    efficiency for disinfection does not vary with pH or in the presence
    of ammonia, and it does not oxidize bromide. As opposed to chlorine,
    which reacts via oxidation and electrophilic substitution, chlorine
    dioxide reacts only by oxidation; this explains why it does not
    produce organochlorine compounds. In addition to this, chlorine
    dioxide is more selective in typical water treatment applications, as
    evidenced by its somewhat lower disinfectant demand as compared with
    chlorine.

         Chlorine dioxide is generally produced by reacting aqueous
    (sodium) chlorite with chlorine (Gordon & Rosenblatt, 1996):

              2ClO2- + HOCl + H+ -> 2ClO2(aq) + Cl- + H2O

         However, under conditions of low initial reactant concentrations
    or in the presence of excess chlorine, the reactant produces chlorate
    ion:

              ClO2- + HOCl -> ClO3- + Cl- + H+

         This reaction scenario is common in generators that
    overchlorinate to achieve high reaction yields based on chlorite ion
    consumption.

         An alternative approach to chlorine dioxide generation is with
    hydrochloric acid (HCl), a process that results in less chlorate
    during production:

              5NaClO2 + 4HCl -> 4ClO2 + 5NaCl + 2H2O

    Chlorite ion is also produced when chlorine dioxide reacts with
    organics (Gordon & Rosenblatt, 1996):

              ClO2 + NOM -> Products + ClO2-

         Chlorine dioxide can also undergo a series of photochemically
    initiated reactions resulting in the formation of chlorate ion (Gordon
    et al., 1995).

         While bromide is not generally oxidized by chlorine dioxide,
    bromate can be formed in the presence of sunlight over a wide range of
    pH values (Gordon & Emmert, 1996). Utilities need to be concerned with
    bromate ion in the chlorine dioxide treatment of drinking-water if the
    water contains bromide and is exposed to sunlight. Practically, this
    means minimizing exposure to sunlight when chlorine dioxide is applied
    in the presence of bromide ion. There appears to be a problem with
    chlorine dioxide producing odour-causing compounds at the tap. This
    has been linked to chlorine dioxide reacting with volatile organic
    compounds derived from new carpets and office products (Hoehn et al.,
    1990).

         Hoigne & Bader (1994) described the kinetics of reaction between
    chlorine dioxide and a wide range of organic and inorganic compounds
    that are of concern in water treatment. Measured rate constants were
    high for nitrite, hydrogen peroxide, ozone, iodide, iron(II), phenolic
    compounds, tertiary amines and thiols. Bromide, ammonia, structures
    containing olefinic double bonds, aromatic hydrocarbons, primary and
    secondary amines, aldehydes, ketones and carbohydrates are unreactive
    under the conditions of water treatment. Chlorine dioxide rapidly
    oxidizes substituted phenoxide anions and many phenols, and
    second-order rate constants have been measured (Rav-Acha & Choshen,
    1987).

    2.4.3  Chloramine reactions

         Chloramination of drinking-water produces THMs (if chloramine is
    formed by chlorination followed by ammonia addition), HAAs, chloral
    hydrate, hydrazine, cyanogen compounds, nitrate, nitrite, organic
    chloramines and 1,1-dichloropropanone (1,1-DCPN) (Dlyamandoglu &
    Selleck, 1992; Kirmeyer et al., 1993, 1995).

         In the presence of even small quantities of organic nitrogen, it
    is possible for chloramination to produce organic chloramines. Several
    researchers have shown that monochloramine readily transfers its
    chlorine at a comparatively rapid rate to organic amines to form
    organohalogen amines (Isaac & Morris, 1983; Bercz & Bawa, 1986).
    Monochloramine was shown to cause binding of radiolabelled halogen to
    nutrients such as tyrosine and folic acid; the amount of binding
    varied with pH but was generally less at neutral pH than at higher pH
    (Bercz & Bawa, 1986). Organic chloramines are much weaker
    disinfectants than inorganic monochloramine but are indistinguishable
    by common analytical methods. Organic chloramine formation may

    necessitate changing chloramination conditions (e.g., ammonia and
    chlorine addition order, chlorine-to-ammonia ratios and contact time).

         HANs and non-halogenated acetonitriles are produced when
    chloramines are reacted with humic materials and amino acids (Trehy et
    al., 1986). The reaction pathway for these products is quite
    complicated and very similar to that for chlorine, with many
    intermediates and by-products formed. In the case of aspartic acid, De
    Leer et al. (1986) demonstrated the presence of at least 11 other
    significant products.

    2.4.4  Ozone reactions

         Ozone has been shown to oxidize bromide to hypobromite and
    bromate, and hypochlorite to chlorate (Glaze et al., 1993; Siddiqui et
    al., 1995; Siddiqui, 1996).

         Bromate generally forms through a combination of molecular ozone
    attack and reaction of bromide with free radical species. The
    molecular ozone mechanism does not account for hydroxyl radicals
    always formed as secondary oxidants from decomposed ozone during water
    treatment. Siddiqui et al. (1995) indicated that there is a radical
    pathway that is influenced by both pH and alkalinity. The hydroxyl
    radical and, to a lesser degree, the carbonate radical (CO3) pathway
    may be more important than the molecular ozone pathway. Oxidants such
    as hydroxyl and carbonate radicals may interact with intermediate
    bromine species, leading to the formation of hypobromite radicals
    (BrO), which eventually undergo disproportionation to form
    hypobromite and bromite (BrO2-). Bromate is then formed through
    oxidation of bromite by ozone. The radical mechanism for the formation
    of bromate includes two decisive reaction steps still involving
    molecular ozone: the formation of hypobromite and oxidation of
    bromite.

         Bromate ion formed through reactions with molecular ozone
    contributes in the range of 30-80% to the overall bromate ion
    formation in NOM-containing waters (von Gunten and Hoigne, 1994).
    Siddiqui et al. (1995) reported up to 65% and 100% bromate ion
    formation through the radical pathway in NOM-free and NOM-containing
    waters, respectively. Differences in NOM-containing waters can be
    attributed to differences in the characteristics of the NOM present. A
    change in mechanism as a function of pH and the competitive roles of
    the free radical (one electron transfer) mechanism above pH 7 versus
    oxygen atom (two electron transfer) mechanism help explain both the
    large variations in bromate ion yield and the sensitivity to reactor
    design, concentration of organic precursors and ozone/bromide ion
    concentrations (Gordon, 1993).

         The presence of bromide ion in a source water further complicates
    the reaction of ozone and leads to the formation of additional DBPs,
    such as bromoform, dibromoacetonitrile (DBAN) and dibromoacetone
    (DBAC) (Siddiqui, 1992; Amy et al., 1993, 1994).

    2.5  Formation of organohalogen disinfectant by-products

         Table 3 summarizes the DBPs identified as being formed from the
    use of chlorine, chlorine dioxide, chloramine and ozone.

         The formation of organochlorine and organobromine compounds
    during drinking-water treatment is a cause of health concern in many
    countries. These compounds include THMs, HAAs, HANs, chloral hydrate,
    chloropicrin, acetohalides, halogenated furanones and other compounds.

    2.5.1  Chlorine organohalogen by-products

         Table 4 summarizes the range of concentrations of chlorinated
    DBPs formed from the reaction of chlorine with NOM, from various
    sources.

         The major chlorination DBPs identified are THMs, HAAs, HANs,
    haloketones, chloropicrin and chloral hydrate. HAAs represent a major
    portion of the non-THM halogenated organic compounds (Miller & Uden,
    1983; Reckhow & Singer, 1985). Many researchers have identified HANs
    and haloketones as other important DBPs (Trehy & Bieber, 1981; Miller
    & Uden, 1983; Oliver, 1983; Reckhow & Singer, 1985). According to an
    AWWARF (1991) study, for all eight utilities tested,
    1,1,1-trichloropropanone (1,1,1-TCPN) was the more prevalent of the
    two measured haloketone compounds. In addition, Kronberg et al. (1988)
    identified the extremely mutagenic compound, MX.

         Despite the fact that HAA formation and THM formation have very
    different pH dependencies, HAA formation correlates strongly with THM
    formation when treatment conditions are relatively uniform and when
    the water has a low bromide concentration (Singer, 1993). DBP
    formation and requisite chlorine dosage for disinfection strongly
    correspond to the concentration of TOC at the point of chlorine
    addition, suggesting that optimized or enhanced removal of organic
    carbon prior to chlorination will decrease the formation of DBPs.

         HAA formation can be appreciable when drinking-water is
    chlorinated under conditions of slightly acidic pH and low bromide
    concentrations. The concentrations of DCA and TCA are similar to the
    concentrations of chloroform, and the total HAA concentration can be
    as much as 50% greater than the THM concentration in the finished
    water on a weight basis.

         McGuire & Meadow (1988) reported that the national average THM
    concentration in the USA was 42 g/litre for drinking-water utilities
    serving more than 100 000 persons, and only 3% of systems were above
    the US maximum contaminant level of 100 g/litre. Amy et al. (1993)
    estimated that the national average THM concentration in the USA was
    40 g/litre, with an average TOC concentration of 3.0 mg/litre. The
    median annual average THM concentration found for utilities among the
    American Water Works Association's (AWWA) Water Industry Database was
    35 g/litre, as compared with 50 g/litre for the non-database
    utilities (Montgomery Watson, Inc., 1993). 


        Table 3. Disinfectant by-products present in disinfected waters

                                                                                                           
    Disinfectant           Significant                  Significant             Significant 
                           organo-halogen               inorganic               non-halogenated
                           products                     products                products
                                                                                                           

    Chlorine/              THMs, HAAs, HANs,            Chlorate (mostly        Aldehydes, cyanoalkanoic 
    hypochlorous acid      chloral hydrate,             from hypochlorite       acids, alkanoic acids, 
                           chloropicrin,                use)                    benzene, carboxylic acids
                           chlorophenols, 
                           N-chloramines, 
                           halofuranones, 
                           bromohydrins

    Chlorine dioxide                                    chlorite, chlorate      unknown

    Chloramine             HANs, cyanogen chloride,     nitrate, nitrite,       aldehydes, ketones
                           organic chloramines,         chlorate, hydrazine
                           chloramino acids, 
                           chloral hydrate, 
                           haloketones

    Ozone                  bromoform, MBA, DBA,         chlorate, iodate,       aldehydes, ketoacids, 
                           DBAC, cyanogen bromide       bromate, hydrogen       ketones, carboxylic 
                                                        peroxide, hypobromous   acids
                                                        acid, epoxides, 
                                                        ozonates
                                                                                                           

    Table 4. Concentration range of chlorinated disinfectant by-products in drinking-watera

                                                                                                                            

    DBPs                  Peters et al. (1990);    Krasner et al.     Nieminski et al.    Koch et al.      Reckhow et al.
                          Peters (1991)            (1989)             (1993)              (1991)           (1990)
                                                                                                                            

    THMs                  3.1-49.5                 30.0-44.0          17.0-51.0           49.0-81.0        201-1280

    HAAs                  <0.5-14.7                13.0-21.0          5.0-25.0            22.0-32.0        118-1230

    HANs                  0.04-1.05                2.5-4.0            0.5-5.0             2.0-2.6          3.0-12.0

    Haloketones           -                        0.9-1.8            0.2-1.6             1.0-2.0          4.8-25.3

    Chlorophenols         -                        -                  0.5-1.0             -                -

    Chloral hydrate       -                        1.7-3.0            -                   -                -

    Chloropicrin          -                        0.1-0.16           <0.1-0.6            -                -

    TOC                   1.7-5.6                  2.9-3.2            1.5-6.0             2.5-3.0          4.8-26.6

    Bromide               100-500                  70-100             -                   170-420          -
                                                                                                                            

    a  All values shown in g/litre, except TOC (mg/litre).
    

         In Germany, 10% of the utilities produced disinfected
    drinking-water with a THM concentration above 10 g/litre; the median
    annual average concentration was between 1 and 4 g/litre, depending
    on raw water quality and size of facility (Haberer, 1994).

         Total THM levels in treated drinking-water were reported in one
    survey in the United Kingdom (Water Research Centre, 1980):
    chlorinated water derived from a lowland river contained a mean level
    of 89.2 g/litre, and that from an upland reservoir, 18.7 g/litre.
    The study also showed that chlorinated groundwater was contaminated by
    THMs to a significantly lesser extent than chlorinated surface waters.

         In a national survey of the water supplies of 70 communities
    serving about 38% of the population in Canada, conducted in the winter
    of 1976-1977, chloroform concentrations in treated water of the
    distribution system 0.8 km from the treatment plant, determined by the
    gas sparge technique, averaged 22.7 g/litre. Levels of the other THMs
    were considerably lower, averaging 2.9 g/litre for
    bromodichloromethane (BDCM), 0.4 g/litre for dibromochloromethane
    (DBCM) and 0.1 g/litre for bromoform. Using direct aqueous injection
    techniques, average concentrations of most of the THMs were higher
    (Health Canada, 1993).

         Samples collected from the distribution systems of eight major
    cities in Saudi Arabia showed that THMs occurred in all the water
    supplies, at concentrations ranging between 0.03 and 41.7 g/litre.
    Median total THM concentrations in several cities were higher during
    the summer than during the winter. In addition, THM concentrations
    were low in cities that did not mix groundwater and desalinated water.
    Brominated THMs dominated (with bromoform the most abundant) and
    existed at the highest concentration levels, whereas chloroform was
    the least prevalent compound. This is the opposite of the occurrence
    pattern found in almost all water distribution systems worldwide
    (Fayad, 1993).

         The concentrations of chloral hydrate in drinking-water in the
    USA were summarized by IARC (1995) and varied from 0.01 to
    28 g/litre. The highest values were found in drinking-water prepared
    from surface water.

         Chlorination of water as well as the combination of ozonation and
    chlorination can lead to the formation of chloropicrin (Merlet et al.,
    1985). In a study conducted for over 25 utilities, very low levels of
    chloropicrin were observed, and chlorination produced maximum
    concentrations of less than 2 g/litre (AWWARF, 1991). The
    chloropicrin appeared to form slowly during the incubation period,
    with concentrations tending to level off at approximately 40 h.

         Dichloroacetonitrile (DCAN) is by far the most predominant HAN
    species detected in water sources with bromide levels of 20 g/litre
    or less. For sources with higher bromide levels (50-80 g/litre),
    bromochloroacetonitrile (BCAN) was the second most prevalent compound.
    However, none of these sources had a DBAN concentration exceeding 0.5
    g/litre, including one source water that had a much higher bromide

    level, 170 g/litre. Thus, it appears that ambient bromide
    concentration is not the only factor influencing the speciation of HAN
    compounds.

         Chlorine can react with phenols to produce mono-, di- or
    trichlorophenols, which can impart tastes and odours to waters. The
    control of chlorophenolic tastes and odours produced when phenol-laden
    water is treated with chlorine is essential. The sources of phenolic
    compounds in water supplies are reported to be industrial wastes.

         In natural waters, one of the most important sources of organic
    nitrogen is proteins and their hydrolysis products. The reaction of
    aqueous chlorine or monochloramine with organic nitrogen may form
    complex organic chloramines (Feng, 1966; Morris et al., 1980; Snyder &
    Margerum, 1982). The formation of  N-chloramines resulting from the
    reaction of amines and chlorine has been reported (Weil & Morris,
    1949; Gray et al., 1979; Morris et al., 1980). Likewise, the
    chlorination of amides has been reported (Morris et al., 1980).

         Nieminski et al. (1993) reported the occurrence of DBPs for Utah
    (USA) water treatment plants. All plants used chlorine for primary and
    secondary disinfection purposes. Overall, THMs and HAAs represented
    75% of the total specific DBPs analysed for the survey; however, total
    DBPs represented only 25-50% of the TOX concentration. THMs
    constituted 64% of the total DBPs by weight; HAAs were 30% of the
    total DBPs by weight and approximately one-half of the total THM
    concentrations. (However, in some waters, HAA concentrations may
    approach or possibly exceed THM concentrations.) HANs, haloketones,
    chlorophenols and chloropicrin represented 3%, 1.5%, 1.0% and 0.5%,
    respectively, of the total surveyed DBPs.

         The occurrence of DBPs in drinking-waters in the USA was
    evaluated at 35 water treatment facilities that had a broad range of
    source water qualities and treatment processes (Krasner et al., 1989).
    THMs were the largest class of DBPs, and HAAs were the next most
    significant class. Aldehydes, by-products of ozonation, were also
    produced by chlorination. Over four quarterly sampling periods, median
    total THM concentrations ranged from 30 to 44 g/litre, with
    chloroform, BDCM, DBCM and bromoform ranges of 9.6-15, 4.1-10, 2.6-4.5
    and 0.33-0.88 g/litre, respectively. Median total HAA concentrations
    ranged from 13 to 21 g/litre, with TCA, DCA, monochloroacetic acid
    (MCA), dibromoacetic acid (DBA) and monobromoacetic acid (MBA) ranges
    of 4.0-6.0, 5.0-7.3, <1-1.2, 0.9-1.5 and <0.5 g/litre,
    respectively.

         Concentrations of DCA and TCA measured in various water sources
    have been summarized by IARC (1995): in Japan, chlorinated
    drinking-water contained 4.5 and 7.5 g of DCA and TCA per litre,
    respectively; rainwater in Germany contained 1.35 g of DCA per litre
    and 0.1-20 g of TCA per litre, whereas groundwater contained 0.05 g
    of TCA per litre; in Australia, a maximum concentration of 200
    g/litre was found for DCA and TCA in chlorinated treated water; and
    chlorinated water in Switzerland contained 3.0 g of TCA per litre.

         In a survey of 20 drinking-waters prepared from different source
    waters in the Netherlands, HAAs were found in all drinking-waters
    prepared from surface water, whereas they could not be detected in
    drinking-waters prepared from groundwaters. Brominated acetic acids
    accounted for 65% of the total acid concentration (Peters et al.,
    1991). In another survey of Dutch drinking waters, the average
    concentration of dihaloacetonitriles was about 5% of the average THM
    concentration (Peters, 1990).

    2.5.2  Chloramine organohalogen by-products

         Chloramine treatment practice involves three potential
    approaches: free chlorine followed by ammonia addition, ammonia
    addition followed by chlorine addition  (in situ production) and
    pre-formed (off-line formation) chloramines. Generally, the objective
    is monochloramine formation. Chlorine followed by ammonia is a common
    approach, and, during the free-chlorine period, DBP formation may
    mimic that of chlorine. Chloramination results in the production of
    THMs (predominantly formed by chlorination followed by ammonia
    addition), HAAs, chloral hydrate, hydrazine, cyanogen compounds,
    organic chloramines and 1,1-DCPN (Dlyamandoglu & Selleck, 1992;
    Singer, 1993; Kirmeyer et al., 1993, 1995). Chloramination
    significantly reduces but does not eliminate THM formation; cyanogen
    chloride and TOX represent the major DBP issues with respect to
    chloramines.

         Scully et al. (1990) identified chloramino acids such as
     N-chloroglycine,  N-chloroleucine and  N-chlorophenylalanine as
    by-products after chlorination of water containing nitrogen or after
    chloramination.

    2.5.3  Chlorine dioxide organohalogen by-products

         Chlorine-free chlorine dioxide does not form THMs (Noack & Doerr,
    1978; Symons et al., 1981). Several studies show that the TOX formed
    with chlorine dioxide is 1-15% of the TOX formed with chlorine under
    the same reaction conditions (Chow & Roberts, 1981; Symons et al.,
    1981; Fleischacker & Randtke, 1983).

         Treatment of phenol-laden source waters with chlorine dioxide
    does not produce the typical chlorophenolic taste and odour compounds
    that are produced when the water is treated using chlorine and is
    effective in removing existing tastes and odours of this type.

    2.5.4  Ozone organohalogen by-products

         Ozonation of drinking-water containing bromide ion has been shown
    to produce hypobromous acid/hypobromite, with hypobromite ion serving
    as an intermediate to bromate formation. In the presence of NOM,
    hypobromous acid produces a host of brominated organic compounds, such
    as bromoform, MBA, DBA, DBAN, cyanogen bromide and DBAC (Glaze et al.,
    1993; Siddiqui & Amy, 1993). Cavanagh et al. (1992) and Glaze et al.
    (1993) reported the identification of bromohydrins, a new group of
    labile brominated organic compounds from the ozonation of a natural

    water in the presence of enhanced levels of bromide. However, results
    by Kristiansen et al. (1994) strongly suggest that the bromohydrins,
    such as 3-bromo-2-methyl-2-butanol, in extracts of unquenched
    disinfected water are artefacts formed from the reaction of excessive
    hypobromous acid with traces of olefins in the extraction solvents and
    not novel DBPs.

         Table 5 compares the median concentrations of various DBPs after
    ozonation and chlorination.

    2.6  Formation of inorganic disinfectant by-products

         Although organic DBPs have been the subject of study over a
    longer time frame, the formation of many inorganic by-products is
    coming under increasing scrutiny. 

    2.6.1  Chlorine inorganic by-products

         Chlorite and chlorate are inorganic by-products formed in some
    chlorine solutions. This is of interest because many small
    drinking-water utilities use hypochlorite solutions as a source of
    free chlorine for disinfection. Bolyard & Fair (1992) examined the
    occurrence of chlorate in samples of untreated source water,
    drinking-water and hypochlorite solutions from 14 sites that use
    hypochlorite solutions. The hypochlorite solutions used were found to
    contain significant levels of chlorate. Chlorite and bromate were also
    found in hypochlorite solutions from these same water utilities.
    Chlorate was present in drinking-water, either as a manufacturing
    by-product or from decomposition reactions occurring during storage.
    Approximately 0.2 mg of chlorate per litre was observed in water
    following the addition of chlorine as sodium hypochlorite at a dose
    sufficient to maintain a residual of 0.45 mg/litre (Andrews &
    Ferguson, 1995). The concentration of chlorite in commercial bleach
    solutions typically ranges from 0.002 to 0.0046 mol/litre; similarly,
    the chlorate concentration ranges from about 0.02 to 0.08 mol/litre
    (Gordon et al., 1995).

         A detailed study by Bolyard & Fair (1992) demonstrated that
    hypochlorite solutions used to disinfect drinking-water contain
    significant levels of chlorite and chlorate. The concentration of
    chlorite ranged from <2 to 130 mg/litre for free available chlorine
    concentrations ranging from 3 to 110 g/litre. The concentration of
    chlorate varied over the range 0.19-50 g/litre, with a median of 12
    g/litre. These solutions also contained bromate levels ranging from
    <2 to 51 mg/litre. The concentrations of chlorate in treated source
    waters ranged from 11 to 660 g/litre. In another study involving 25
    samples from plants using gaseous chlorine, no chlorate was detected,
    indicating that the use of gaseous chlorine does not produce chlorate
    (Bolyard & Fair, 1992). Nieminski et al. (1993) measured chlorate and
    chlorite for six water treatment plants that use liquid chlorine
    (i.e., hypochlorite) and found chlorate concentrations ranging from 40
    to 700 g/litre, with no chlorite or bromate detected in finished
    waters. These chlorate concentrations may be attributed to high

    Table 5. Median concentrations of organic disinfectant by-products 
             in drinking-water

                                                                             
    DBPs                Median concentration          Median concentration 
                        (g/litre): chlorinationa     (g/litre): ozonationb
                                                                             

    THMs                          40                         <1.0
      Chloroform                  15                            -
      BDCM                        10                            -
      DBCM                       4.5                            -
      Bromoform                 0.57                         <1.0

    HANs                         2.5                         <1.0
      TCAN                    <0.012                            -
      DCAN                       1.1                            -
      BCAN                      0.58                            -
      DBAN                      0.48                         <1.0

    Haloketones                 0.94                            -
      DCPN                      0.46                            -
      TCPN                      0.35                            -

    HAAs                          20                         <5.0
      MCA                        1.2                            -
      DCA                        6.8                            -
      TCA                        5.8                            -
      MBA                       <0.5                         <1.0
      DBA                        1.5                         <5.0

    Aldehydes                    7.8                           45
      Formaldehyde               5.1                           20
      Acetaldehyde               2.7                           11
      Glyoxal                      -                            9
      Methylglyoxal                -                            5

    Chloral hydrate              3.0                            -

    Ketoacids                      -                           75

    Trichlorophenol             <0.4                            -
                                                                             

    a Krasner et al. (1989).
    b Siddiqui et al. (1993).

    concentrations of chlorate, ranging from 1000 to 8000 mg/litre,
    detected in a bleach used for disinfection and resulting from the
    decomposition of hypochlorite stock solution. However, no chlorite or
    chlorate was detected in any of the samples of finished water of the
    treatment plants that apply gaseous chlorine. Chlorate formation is
    expected to be minimal in low-strength hypochlorite solutions freshly
    prepared from calcium hypochlorite, because of the low hypochlorite
    concentration and only mildly alkaline pH.

    2.6.2  Chloramine inorganic by-products

         Inorganic by-products of chloramination include nitrate, nitrite,
    hydrazine and, to some extent, chlorate (Dlyamandoglu & Selleck, 1992;
    Kirmeyer et al., 1995). 

    2.6.3  Chlorine dioxide inorganic by-products

         The major inorganic by-products of chlorine dioxide disinfection
    have been identified as chlorite and chlorate. Andrews & Ferguson
    (1995) measured a chlorate concentration of 0.38 mg/litre when a
    chlorine dioxide residual of 0.33 mg/litre was maintained. The
    application of chlorine dioxide produces about 0.5-0.7 mg of chlorite
    and 0.3 mg of chlorate per mg of chlorine dioxide consumed or applied
    (Andrews & Ferguson, 1995).

    2.6.4  Ozone inorganic by-products

         When bromide or iodide ions are present in waters, some of the
    halogen-containing oxidants that can be produced during ozonation
    include free bromine, hypobromous acid, hypobromite ion, bromate ion,
    free iodine, hypoiodous acid and iodate ion.

         During the oxidation or chemical disinfection of natural waters
    containing bromide ion with ozone, bromate ion can be formed at
    concentrations ranging from 0 to 150 g/litre under normal water
    treatment conditions (Siddiqui, 1992). Chlorate formation with an
    initial total chlorine concentration of 0.6 mg/litre was evaluated at
    pH levels of 8.0, 7.0 and 6.0, and chlorate concentrations ranging
    from 10 to 106 g/litre were formed after ozonation (Siddiqui et al.,
    1996a).

         It has been reported that ozone reacts with many metal ions and
    with cyanide ion (Hoigne et al., 1985; Yang & Neely, 1986). Bailey
    (1978) discussed the formation of ozonates, compounds of metal cations
    having the general formula M+O3-. Hydrogen peroxide has been
    identified as a by-product of ozonation of organic unsaturated
    compounds (Bailey, 1978).

         Table 6 provides the range of bromate concentrations normally
    encountered in drinking-waters with a variety of source water
    characteristics after ozonation.


        Table 6. Summary of bromate ion formation potentials in different source waters under different conditions following 
             ozonation

                                                                                                                               

    Na      Bromide        Ozone          pH           Alkalinity     DOC           Bromate         Reference
            (g/litre)     (mg/litre)                  (mg/litre)     (mg/litre)    (g/litre)
                                                                                                                               

    18      10-800         1-9.3          5.6-9.4      20-132         2.2-8.2       <5-60           Krasner et al. (1992)

    4       60-340         3-12           6.5-8.5      90-230         3-7           <5-40           Siddiqui & Amy (1993)

    28      10-100         2-4            6.8-8.8      20-120         0.3-11        <5-100          Amy et al. (1993, 1994)

    4       12-37          0-3.97         7.8          N/A            N/A           <7-35           Hautman & Bolyard (1993)

    1       500            2.3-9.5        7.2-8.3      N/A            N/A           13-293          Yamada (1993)

    23      12-207         0.3-4.3        5.7-8.2      14-246         0.5-6.8       <2-16           Legube et al. (1993)

    8       107-237        1-5            6.8-8.0      N/A            2-5           <5-50           Kruithof & Meijers (1993)
                                                                                                                               

    a  N = number of sources studied.
    

    2.7  Formation of non-halogenated organic disinfectant by-products

    2.7.1  Chlorine organic by-products

         Lykins & Clark (1988) conducted a 1-year pilot plant study of the
    effects of ozone and chlorine and determined that the concentration of
    aldehydes increased by 144% upon ozonation. In the chlorinated stream,
    the concentration of these aldehydes increased by 56%. This study
    indicates that aldehyde formation, although greater with ozone, is not
    unique to ozonation, but is associated with chlorination and other
    oxidants as well.

    2.7.2  Chloramine organic by-products

         When Suwannee River (USA) fulvic acid was reacted with aqueous
    solutions of 15N-labelled chloramine and 15N-labelled ammonia,
    lyophilized products exhibiting nuclear magnetic resonances between 90
    and 120 ppm were observed, denoting the formation of amides,
    enaminones and aminoquinones (Ginwalla & Mikita, 1992). This
    represents evidence for the formation of nitrogen-containing compounds
    from the chloramination of NOM in natural waters.

         Amino acids, peptides and amino sugars were chlorinated under
    various chlorine/nitrogen ratios (Bruchet et al., 1992). Six natural
    amino acids (alanine, methionine, valine, phenylalanine, leucine and
    isoleucine) were shown to induce tastes and odours at concentrations
    in the range of 10-20 g/litre. Detectable odours were consistently
    induced in a multicomponent mixture containing each of these amino
    acids after a 2-h contact time with chlorine. Investigation of the
    by-products indicated that the odours generated were systematically
    linked to the aliphatic aldehydes formed. The peptides investigated
    had varying degrees of odour formation potential, while the amino
    sugars did not impart any odour. Chlorinous odours occasionally
    detected during these experiments were found to be due to organic
    chloramines and other oxidation by-products.

    2.7.3  Chlorine dioxide organic by-products

         Gilli (1990) showed the formation of carbonyl compounds
    (34 g/litre) such as  n-valeraldehyde (7-15 g/litre), formaldehyde
    (3.4-9 g/litre), acetaldehyde (4.5 g/litre) and acetone (3.2
    g/litre) after using chlorine dioxide.

    2.7.4  Ozone organic by-products

         Ozone aliphatic oxidation products from organic impurities in
    water are usually acids, ketones, aldehydes and alcohols. So-called
    ultimate oxidation products of organic materials are carbon dioxide,
    water, oxalic acid and acetic acid. However, ozonation conditions
    generally employed in treating drinking-water are rarely sufficient to
    form high concentrations of these ultimate products.

         When source waters containing NOM and unsaturated organic
    compounds are ozonated, ozonides, peroxides, diperoxides, triperoxides
    and peroxy acids, for example, can be produced. The limited research
    that has been conducted in aqueous solutions indicates that these
    intermediates decompose readily in water to form products such as
    aldehydes, ketones, carboxylic acids and ketoacids.

         Coleman et al. (1992) identified numerous compounds in addition
    to the following in ozonated humic samples: monocarboxylic acids up to
    C-24, dicarboxylic acids up to C-10, ketoacids, furan carboxylic
    acids, and benzene mono-, di- and tricarboxylic acids. Among the
    various aldehydes, Paode et al. (1997) found four (formaldehyde,
    acetaldehyde, glyoxal and methylglyoxal) to be dominant. Table 7
    provides a range of concentrations for aldehydes from the ozonation of
    a variety of source waters.

    2.8  Influence of source water characteristics on the amount and type
         of by-products produced

         The extensive literature pertaining to DBP levels in disinfected
    source waters and control of DBPs by various treatment processes
    attests to the wide variety of factors influencing DBP formation and
    the complex interrelationships between these factors. Variables
    including the concentration and characteristics of precursor material,
    pH, chlorine concentration, bromide level, presence of
    chlorine-demanding substances such as ammonia, temperature and contact
    time all play a role in DBP formation reactions.

    2.8.1  Effect of natural organic matter and UV absorbance at 254 nm

         NOM consists of a mixture of humic substances (humic and fulvic
    acids) and non-humic (hydrophilic) material. Both the amount (as
    indicated by TOC or UVA254) and the character (as described by UVA254)
    of NOM can affect DBP formation. NOM provides the precursor material
    from which organic DBPs are formed; consequently, increasing
    concentrations of NOM lead to increasing concentrations of
    by-products. This relationship has led to the use of TOC and UVA254
    measurements as surrogate parameters for estimating the extent of DBP
    formation. 

         The removal of NOM is strongly influenced by those properties
    embodying the size, structure and functionality of this heterogeneous
    mixture. The humic acids are more reactive than fulvic acids with
    chlorine (Reckhow et al., 1990) and ozone, in terms of both
    oxidant/disinfectant demand and DBP formation. Processes such as
    coagulation, adsorption and membrane filtration are separation
    processes that remove NOM intact, while ozonation transforms part of
    the NOM into biodegradable organic matter, potentially removable by
    biofiltration. Coagulation preferentially removes humic/higher
    molecular weight NOM; the selectivity of membranes for NOM removal is
    largely dictated by the molecular weight cutoff of the membrane; the
    use of granular activated carbon (GAC) requires a significant empty
    bed contact time; biofiltration can remove only the rapidly
    biodegradable NOM fraction.


        Table 7. Effect of ozone dose and TOC on non-halogenated organic by-products

                                                                                                                       
    Ozone dose     TOC           Formal       Acetal        Glyoxal       Methyl-glyoxal    Reference
    (mg/litre)     (mg/litre)    (g/litre)   (g/litre)    (g/litre)    (g/litre)
                                                                                                                       

    1.2-4.4        2.66          8-24         2-4           4-11          4-15              Miltner et al. (1992)

    1.0-9.2        1.0-25.9      3-30         7-65          3-15          3-35              Weinberg et al. (1993)

    5.5-28.5       5.4-17.4      58-567       6-28          15-166        17-54             Schechter & Singer (1995)
                                                                                                                       
    

         In an investigation of the nature of humic and fulvic acids
    isolated from a variety of natural waters, Reckhow et al. (1990) found
    that the fulvic fractions had a lower aromatic content and smaller
    molecular size than the humic fractions. UV absorbance was
    correspondingly higher for the humic fractions, owing to the higher
    aromatic content and larger size. These researchers also found that
    for all of the organic material investigated, the production of
    chloroform, TCA, DCA and DCAN was higher upon chlorination of the
    humic fractions than upon chlorination of the corresponding fulvic
    fractions. These findings support the findings of other researchers
    and show that the UV absorbance measurement is an indicator of the
    nature of the precursor material present in a sample. This
    measurement, in conjunction with the TOC (or DOC) measurement, can be
    employed in the evaluation of data to provide an indication of the
    reactivity of NOM towards forming DBPs.

         The reaction of ozone with NOM can occur directly or by radical
    processes. The disappearance of disinfecting chemical is influenced by
    the type and concentration of NOM present in natural waters. Direct
    consumption of these chemicals is greater when the UV absorbance (due
    to electrophilic and nucleophilic sites of NOM) of the source water is
    significant, resulting in decreased DBP formation potential.

         It appears that the nature of the organic material in a source
    water may have some impact on the relative concentrations of THMs and
    HAAs formed upon chlorination. Treatment techniques that lower the
    levels of DOC without affecting bromide levels have been implicated in
    a shift from chlorinated to brominated THM compounds. This is of
    concern because the theoretical risk to humans varies for the
    individual THMs, with the brominated species generally being of more
    concern (Bull & Kopfler, 1991).

    2.8.2  Effect of pH

         The impact of pH on THM concentrations has been reported by a
    number of researchers since THMs in drinking-water first came to the
    attention of the water industry (Stevens et al., 1976; Lange &
    Kawczynski, 1978; Trussell & Umphres, 1978). More recently, the impact
    of pH on a number of other chlorination by-products has been reported
    (Miller & Uden, 1983; Reckhow & Singer, 1985). The rate of THM
    formation increases with the pH (Stevens et al., 1976). Kavanaugh et
    al. (1980) reported a 3-fold increase in the reaction rate per unit
    pH.

         In general, increasing pH has been associated with increasing
    concentrations of THMs and decreasing concentrations of HAAs (pH
    primarily impacting TCA), HANs and haloketones. The concentrations of
    TCA tend to be higher in waters with pH levels less than 8.0 than in
    waters with pH levels greater than 8.0; a less marked trend is
    observed for DCA. Other researchers have reported similar findings
    with respect to the pH dependency of HAA concentrations. For example,
    Stevens et al. (1976) found that TCA concentrations were significantly
    lower at a pH of 9.4 than at pH levels of approximately 5 and 7. TCA
    was by far the most predominant of the measured HAA species at six of

    the eight utilities surveyed. Carlson & Hardy (1998) reported that at
    pH levels greater than 9.0, THM formation decreased with increasing
    pH. It is possible that the shift in chlorine species from
    hypochlorous acid to hypochlorite affects THM formation during short
    reaction times.

         AWWARF (1991) observed no relationship between pH and the
    concentrations of THMs at eight utilities over time, suggesting that
    although THM concentrations for a particular water are known to be pH
    dependent, factors other than pH influence THM concentrations over a
    variety of source waters. Nieminski et al. (1993) reported that
    treatment plants with a pH of about 5.5 in finished water produced
    equal amounts of THMs and HAAs, whereas plants with pHs greater than
    7.0 in finished water produced higher amounts of THMs as compared with
    HAAs.

         No strong relationship has been observed between HAN
    concentration and pH over time. Within the approximate pH range 7-8.5,
    HAN concentrations increased slightly over time. In general, a trend
    of decreasing HAN concentrations with increasing pH would be expected,
    since these compounds are known to undergo base-catalysed hydrolysis
    and have been identified as intermediates in the formation of
    chloroform (Reckhow & Singer, 1985). Therefore, these compounds may be
    unstable in the presence of free chlorine or under basic conditions.
    In general, after an initial formation period, HAN and haloketone
    concentrations level off or begin to decline over the remainder of the
    reaction period. This indicates that base-catalysed hydrolysis may not
    be a significant mechanism of reaction for the relatively low pH
    sources.

         Stevens et al. (1989) evaluated the effects of pH and reaction
    time (4, 48 and 144 h) on the formation of chloral hydrate. Chloral
    hydrate formation increased over time at pH 5 and 7, whereas chloral
    hydrate that had formed within 4 h at pH 9.4 decayed over time at the
    elevated pH.

         The pH of the source water can also affect the formation of
    by-products after chloramine addition. The disproportionation of
    monochloramine, which is an important reaction leading to an oxidant
    loss, has been shown by several researchers to be catalysed by
    hydrogen ion, phosphate, carbonate and silicate (Valentine & Solomon,
    1987).

         Humic acids have shown reaction rates with chlorine dioxide that
    increased by a factor of 3 per pH unit (pH 4-8) (Hoigne & Bader,
    1994).

         In addition to the impact of pH on THM and HAA formation noted
    above, overall TOX formation decreases with increasing pH. Many of the
    halogenated DBPs tend to hydrolyse at alkaline pH levels (>8.0)
    (Singer, 1994a). This has significant implications, for example, for
    precipitative softening facilities.

         pH has a strong effect on aldehyde formation (Schechter & Singer,
    1995). Higher ozonation pH values produced lower amounts of aldehydes,
    supporting the theory that these DBPs are formed primarily through the
    direct molecular ozone reaction pathway, as opposed to the radical
    pathway. These results may also reflect greater destruction of
    aldehydes by hydroxyl radicals at elevated pH levels.

    2.8.3  Effect of bromide

         The presence of bromide ion during water treatment disinfection
    can lead to the formation of DBPs such as brominated organics and
    bromate ion. Low but significant levels of bromide, the ultimate
    precursor to bromate and other brominated compounds, may occur in
    drinking-water sources as a result of pollution and saltwater
    intrusion in addition to bromide from natural sources. An
    understanding of the sources and levels of bromide ion in different
    source waters is crucial for an understanding of the bromate ion
    formation potential in drinking-waters. There are no known treatment
    techniques available for economically removing bromide ion present in
    source waters during drinking-water treatment.

         The impact of bromide on the speciation of DBPs within a class of
    compounds such as THMs or HAAs has been discussed by Cooper et al.
    (1983, 1985) and Amy et al. (1998). Rook et al. (1978) reported that
    bromine is more effective than chlorine in participating in
    substitution reactions with organic molecules; furthermore, precursor
    materials may differ in their susceptibility to bromination versus
    chlorination reactions. Hypobromous acid formed from bromide may also
    react with ammonia to form bromamines (Galal-Gorchev & Morris, 1965).

    2.8.4  Effect of reaction rates

         After chlorine addition, there is a period of rapid THM formation
    for the initial few hours (e.g., 4 h), followed by a decline in the
    rate of THM formation, suggesting fast and slow NOM reactive sites.
    Many authors have indicated that the concentration of chloroform
    appears to increase slowly even after 96 h, suggesting that as long as
    low concentrations of free chlorine are present, chloroform continues
    to form. Bromochlorinated THM species have been found to form more
    rapidly than chloroform. Further data from many sources indicate that
    bromoform formation slows at approximately 7-8 h and levels off almost
    completely after 20 h (AWWARF, 1991; Koch et al., 1991). The same
    general kinetic trend observed for THMs also appears to apply to HAAs.
    A period of rapid formation occurs during the first 4-8 h, followed by
    a reduction in the formation rate. In general, for most sources,
    concentrations of chlorinated HAAs appear to slowly increase even
    after 96 h, while the formation of DBA levels off after about 18-20 h.

         Miller & Uden (1983) observed that nearly 90% of the final
    concentrations of THMs, TCA and DCA form within the first 24 h of
    chlorine addition to waters containing NOM. Reckhow et al. (1990)
    found that although waters containing precursor materials isolated
    from six different water sources differed in their yields of
    chlorinated organic by-products, the formation curves for chloroform,

    TCA and DCA had the same general shapes for all six precursor
    materials. Some researchers have suggested that DCA may be an
    intermediate in TCA formation; however, for all eight source waters
    studied, both DCA and TCA concentrations increased or remained stable
    throughout the 96-h reaction period, suggesting that DCA was an
    end-product (AWWARF, 1991). Carlson & Hardy (1998) indicated that HAA
    formation followed a pattern similar to that of THM formation. As with
    the THMs, HAA formation rate appeared to be rapid for the first 30
    min; after 30 min, HAAs formed at nearly a constant rate in four of
    the source waters studied.

         Different trends were observed in the HAN concentrations of
    different source waters. For two source waters, HAN levels formed
    rapidly for the first 8 h and continued to increase slowly or levelled
    off after 96 h (AWWARF, 1991). DBAN levels remained relatively stable
    over the 96 h, as did BCAN and DCAN levels. For other sources, levels
    of HANs consisting mostly of DCAN increased rapidly up to 4-8 h and
    began to decline by the end of the 96-h period. For these sources,
    BCAN appeared to be slightly more stable than DCAN.

         Very low levels of chloropicrin formation have been observed by
    many researchers (AWWARF, 1991). The highest concentration observed
    was 4.0 g/litre. Chloropicrin appears to form slowly during the
    incubation period, with concentrations tending to level off at
    approximately 40 h.

    2.8.5  Effect of temperature

         The formation rates of THMs, HAAs, bromate ion and HANs have been
    shown to increase with temperature (AWWARF, 1991; Siddiqui & Amy,
    1993). Both haloketone and chloropicrin levels were found to be higher
    at a lower temperature, while the concentrations of other DBP species
    were similar or not significantly different. These results suggest
    that a higher temperature allows for more rapid progression of the
    transformation of haloketones to other by-products. In studies on the
    effect of temperature on THMs, Peters et al. (1980) found an Arrhenius
    dependency between the rate constant and temperature with an
    activation energy of 10-20 kJ/mol.

         The impact of temperature on THMs was strongest at longer contact
    times (Carlson & Hardy, 1998). On a conceptual basis, it may be that
    rapidly forming compounds are more reactive and form DBPs regardless
    of temperature. On the other hand, slowly forming compounds require
    higher activation energy, and an increase in the temperature supplies
    the energy. In addition to reaction kinetics, the temperature of a
    source water can also affect disinfection efficiency. The biocidal
    effectiveness of monochloramine is significantly less than that of
    free chlorine and is dependent on temperature, pH and residual
    concentration.

    2.8.6  Effect of alkalinity

         Although pH is a very influential variable and alkalinity affects
    pH, alkalinity itself does not appear to directly affect the formation
    of THMs and HAAs (by chlorination) and has only a slight effect on
    aldehydes and other organic by-products following ozonation (Andrews
    et al., 1996). However, the majority of studies on the effect of
    alkalinity on the formation of bromate during ozonation indicate that
    increased alkalinity increases bromate formation (Siddiqui et al.,
    1995). The quantity of aldehydes produced remains approximately
    constant for similar changes in alkalinity and pH; however, deviation
    from equivalent changes in pH and alkalinity results in increased
    aldehyde concentrations. Therefore, conditions of high alkalinity and
    low pH or low alkalinity and high pH produce greater quantities of
    aldehydes than do intermediate values of these parameters (Andrews et
    al., 1996).

    2.9  Influence of water treatment variables on the amount and type of
         by-products produced

         Since DBPs are formed by all of the above chemical disinfectants,
    the adoption of alternative disinfectants for DBP control often means
    only a trade-off between one group of DBPs and another. The most
    effective DBP control strategy is organic precursor (NOM) removal
    through enhanced coagulation, biofiltration, GAC or membrane
    filtration. There has been little success with bromide removal. Other
    DBP control options include water quality modifications -- for
    example, acid or ammonia addition for bromate minimization.

    2.9.1  Effect of ammonia

         The presence of ammonia in source waters during disinfection can
    cause chlorine and ozone demand and participation in the formation of
    by-products such as nitrate, cyanogen chloride and other nitrogenous
    compounds.

         Ammonia also does not consume chlorine dioxide. In contrast to
    chlorine, chlorine dioxide can therefore be considered as a virucide
    when ammonia is present. This might be one of the historical reasons
    why chlorine dioxide has been adopted as a disinfectant by some
    treatment plants using well oxidized waters but containing changing
    ammonia concentrations. The addition of ammonia has been shown to
    reduce the formation of bromate after ozonation (Siddiqui et al.,
    1995), and the ammonia has been shown to participate in the formation
    of HANs and cyanogen bromide (CNBr) (Siddiqui & Amy, 1993).

         The growth of nitrifying bacteria is a potential problem in
    chloraminated water supplies or chlorination of sources containing
    nitrogen. In a study conducted by Cunliffe (1991), nitrifying bacteria
    were detected in 64% of samples collected from five chloraminated
    water supplies in South Australia and in 21% of samples that contained
    more than 5 mg of monochloramine per litre. Increased numbers of the
    bacteria were associated with monochloramine decay within the
    distribution systems.

    2.9.2  Effect of disinfectant dose

         Chlorine dose is a factor affecting the type and concentration of
    DBPs formed. The THM level rises with increasing chlorine dose
    (Kavanaugh et al., 1980). However, there is some disagreement
    regarding the quantitative relations between chlorine concentration
    and THM levels (or the rate of THM production). Most investigators
    found a linear relationship between chlorine consumption and THM
    production, with an order of reaction greater than or equal to unity
    (Trussell & Umphres, 1978; Kavanaugh et al., 1980). However, it is
    also possible that the order of reaction changes during the course of
    the reaction.

         Reckhow & Singer (1985) found that the concentration of DBP
    intermediates such as DCAN and 1,1,1-TCPN formed after 72 h of
    reaction time was dependent on chlorine dose. DCAN, which was measured
    at a concentration of approximately 5 g/litre at a chlorine dose of
    10 mg/litre, was not detected in samples dosed with 50 mg of chlorine
    per litre. The concentration of chloroform was about 150 g/litre in a
    sample dosed with 10 mg of chlorine per litre but was approximately
    200 g/litre in a sample dosed with 20 mg of chlorine per litre. Thus,
    it is imperative to have uniform chlorine doses for performing DBP
    formation kinetic measurements.

         Since chloramine residuals are longer-lasting than free chlorine
    residuals, the doses for each set of chlorinated and chloraminated
    samples will be different in order to achieve the prescribed target
    residual. The disappearance of chloramines can be explained
    approximately by a second-order reaction. However, as the chlorine
    dose increased, the observed rate constant was found to decrease, then
    increased after reaching a minimum value (Dlyamandoglu & Selleck,
    1992). Below the chlorine dose at the minimum value of the observed
    rate constant, the rate constant was proportional to the 1.4 power of
    the chlorine dose, regardless of the ammonia concentration (Yamamoto
    et al., 1988).

    2.9.3  Effect of advanced oxidation processes

         Water utilities can add treatment processes that remove DBP
    precursors or DBPs. Many utilities will be using both approaches. The
    hydrogen peroxide/UV process, an advanced oxidation process, offers
    small water utilities a treatment process with the potential to
    provide primary disinfection and a method of DBP control (Symons &
    Worley, 1995). This process has been shown to oxidize dissolved
    organic halogens and decrease TOC. TOC removal as a function of UV
    dose has also been demonstrated by Worley (1994), with TOC removals of
    between 0% and 80%. Andrews et al. (1996) evaluated the effect of the
    hydrogen peroxide/UV process on THM formation and concluded that this
    process is only slightly effective in reducing the formation of DBPs.
    However, using hydrogen peroxide at 1 mg/litre in combination with UV
    effectively reduced or prevented the formation of aldehydes. Other
    advanced oxidation processes (e.g., hydrogen peroxide/ozone, ozone/UV)
    involving hydroxyl (and hydroperoxyl) radical formation may provide
    similar opportunities.

    2.9.4  Effect of chemical coagulation

         Enhanced coagulation and softening will remove TOC. Enhanced
    coagulation is characterized by coagulant doses greater than those
    required for optimum turbidity removal; as an alternative to higher
    doses, a combination of acid (pH depression) and coagulant addition
    can be practised.

         All organic DBPs were reduced by the addition of commonly used
    coagulants. Iron-based coagulants, such as ferric chloride, were
    consistently more effective than alum in removing NOM (Crozes et al.,
    1995). Alum coagulation removed all DBP precursors to a significant
    extent. The percentage removals showed the same trends as, but were
    not identical to, the percentage removals of TOC and UV absorbance. UV
    absorbance was removed to a somewhat greater extent than TOC. Hence,
    TOC and UV absorbance can serve as surrogate parameters for DBP
    formation potential. A fairly good correlation was observed between
    the ratio of HAAs to THMs and the ratio of UV absorbance to TOC,
    indicating that the relative concentrations of HAAs and THMs do to
    some extent depend on the nature of the precursor material. However,
    more data from waters of different qualities would be required to
    evaluate the validity of this relationship.

         The effectiveness of coagulants in removing DBP precursors is
    dependent upon the molecular size of the dissolved organic matter.
    Normally, higher molecular weight fractions are effectively removed
    through coagulation. In a study conducted by Teng & Veenstra (1995),
    water containing dissolved organic matter with molecular weights in
    the range 1000-10 000 daltons generally produced the largest amounts
    of THMs and HAAs under conditions of free chlorination. Coagulation
    and ozonation shift a proportionately greater amount of the THM and
    HAA formation potential to the smallest molecular weight range (<1000
    daltons).

         Coagulation and filtration remove NOM but not bromide, hence
    increasing the ratio of bromide to TOC. As a result, the subsequent
    use of chlorine generally favours the formation of brominated organic
    DBPs.

    2.9.5  Effect of pre-ozonation

         Several studies of ozone oxidation followed by chlorination
    showed increased, rather than decreased, levels of THMs (Trussell &
    Umphres, 1978). This is attributed, at least partially, to the
    formation of aldehydes by ozonation. Another possibility is
    hydroxylation of aromatic compounds to produce  m-dihydroxy aromatic
    derivatives, which are known THM precursors (Lykins & Clark, 1988).
    Although the aldehydes produced contain polar groupings, they are
    nevertheless not easily removed during the flocculation step by
    complexation with aluminium or iron salts. A convenient and more
    appropriate method for the removal of the aldehydes formed during
    ozonation is the incorporation of a biological treatment step
    (biofiltration) in the water treatment process following ozone
    oxidation.

         Pre-ozonation can have both positive and negative effects on DBP
    formation. Pre-ozonation with typical water treatment dosages and
    bicarbonate levels has been shown to remove TCA and DCAN precursors.
    However, such treatment can result in no net change in the DCA
    precursors and may lead to an increase in 1,1,1-TCPN precursors
    (Reckhow & Singer, 1985). According to Teng & Veenstra (1995),
    pre-ozonation resulted in enhanced formation of DCA during
    chlorination and chloramination in the presence of precursors in the
    <1000 dalton molecular weight range. They also indicated that
    pre-ozonation plus chloramination controlled the overall production of
    THMs and HAAs. However, the use of pre-ozonation coupled with free
    chlorination increased the yield of DCA for both the hydrophilic and
    hydrophobic fractions of NOM as compared with free chlorination alone.

         With ozone-chlorine treatment, chloral hydrate formation can be
    enhanced. This behaviour, which has also been observed for DCA,
    suggests that the reaction that produces chloral hydrate is
    accelerated under the conditions of ozonation in combination with
    prechlorination and warm water temperatures (LeBel et al., 1995).

         Ozonation in the presence of traces of hypochlorite ion can form
    inorganic by-products such as chlorate. Siddiqui et al. (1996a) showed
    that if there is any residual chlorine present, ozone can potentially
    oxidize hypochlorite ion to chlorate.

         Coleman et al. (1992) suggested that brominated MX analogues and
    other mixed bromochlorinated by-products formed after ozonation and
    chlorination can possibly increase mutagenic activity.

    2.9.6  Effect of biofiltration

         Biofiltration (ozone-sand filtration or ozone-GAC) can
    potentially reduce TOC, organic by-products and the formation of
    halogenated DBPs.

         Passage of ozonated water samples through a rapid sand filter
    reduced the concentration of aldehydes by 62% (Lykins & Clark, 1988).
    Chlorinated samples experienced a 26% reduction in aldehyde
    concentrations under the same conditions. These reductions in aldehyde
    levels are attributable to biological activity in the sand filters. If
    GAC filtration follows sand filtration, ozone oxidation can be
    expected to promote more bioactivity in the GAC filter, because a
    better colonization environment is provided for microorganisms on GAC
    particles than on sand. Thus, the biological conversion of oxidized
    water impurities to carbon dioxide and water will be greater during
    passage through GAC media. Similar aldehyde removals have been
    observed by several researchers (Van Hoof et al., 1985; Sketchell et
    al., 1995).

         Drinking-water treatment techniques that remove organic
    contaminants without affecting bromide concentrations cause a shift in
    the formation of DBPs towards brominated DBPs. Sketchell et al. (1995)
    studied three sources containing three different DOC levels and
    ambient bromides, which were filtered through biologically active GAC

    filters. Analysis of treated waters showed no removal of bromide ion
    and a shift towards more brominated organo-DBPs. THM levels after
    treatment with GAC with no added ozone decreased from 900-1700 to
    100-700 g/litre. These water sources contained DOC levels ranging
    from 10 to 25 mg/litre and high concentrations of biodegradable DOC
    (DOC removals ranged from 60% to 80% after GAC treatment).

         Table 8 summarizes the effects of ozonation and biofiltration on
    the formation of DBPs from various sources.

    2.10  Comparative assessment of disinfectants

         A comparative assessment (Table 9) of various disinfecting
    chemicals for pre-disinfection (or oxidation) and post-disinfection
    and maintaining a residual for 5 days to simulate concentrations in
    the distribution system showed that the use of free chlorine produces
    the largest concentration of halogenated DBPs (Clark et al., 1994).
    The concentration of DBPs may be reduced by adding ozone or chlorine
    dioxide as a preoxidant, although enhanced formation has been
    observed.

         Table 10 summarizes the effects of water quality and treatment
    variables on the formation of DBPs.

    2.11  Alternative strategies for disinfectant by-product control

         The concern about chlorite, bromate, chlorate and other DBPs in
    drinking-water following treatment with disinfectants has stimulated
    research into ways to eliminate the production or enhance the removal
    of DBPs. Strategies for DBP control include source control, precursor
    removal, use of alternative disinfectants and removal of DBPs by
    technologies such as air stripping, activated carbon, UV light and
    advanced oxidation technologies. For DBPs that can arise in
    hypochlorite solutions (e.g., chlorate), the purity and storage
    conditions of the solutions are important concerns.

    2.11.1  Source control

         Source control options involve controlling nutrient inputs to
    waters (e.g., algae growth control) (Hoehn et al., 1990) that are used
    as drinking-water sources, watershed management (e.g., constructing
    stormwater detention basins), saltwater intrusion control (e.g.,
    development of structural or hydrodynamic barriers to control TOC and
    bromide), and using the concept known as aquifer storage and recovery
    (e.g., drawing water during seasons when the quality of the water is
    best) (Singer, 1994a).

    2.11.2  Organohalogen by-products

         Strategies for control of organohalogen by-products include
    removal of DBPs that are formed using technologies such as oxidation,
    aeration and carbon adsorption (Clark et al., 1994); and removal of
    precursors using treatment techniques such as conventional treatment,


        Table 8. Effects of ozonation and biofiltration on chlorine organic by-products

                                                                                                                   
    DBPs             Ozonation         Biofiltration     Ozonation + biofiltration     Reference
                     (% change)        (% change)        (% change)
                                                                                                                   

    THMs             -20               -20               -40 (chlorine)                Speitel et al. (1993)

    HAAs             -10               -13               -25 (chlorine)                Speitel et al. (1993)

    Chloropicrin     +50 to +250                         -50 to -100 (chlorine)        Miltner et al. (1992)

    Aldehydes        +425 to +1300     -40 to -50        -92 to -98                    Miltner et al. (1992)

    TOX              -30                                 -51 (chlorine)                Miltner et al. (1992)

    TOX              -10               -38               -47 (chlorine)                Shukairy & Summers (1992)

    TOX              +32               -69               -60 (monochloramine)          Shukairy & Summers (1992)
                                                                                                                   

    Table 9. Comparative assessment of organic disinfectant by-products (g/litre) in distribution systemsa,b

                                                                                                                  
    DBPs              Sand-Cl2    Cl2-Sand-Cl2   O3-Sand-Cl2   NH2Cl-Sand-NH2Cl   O3-Sand-NH2Cl   ClO2-Sand-Cl2
                                                                                                                  

    THMs                236.0         225.0         154.0             9.0               3.2           138.0

    HAAs                 60.0         146.0          82.0            14.0               9.0            44.0

    HANs                  3.1           2.9           2.7            <0.1              <0.1            <0.1

    Haloketones           2.1           2.6           2.6            <0.1              <0.1             4.2

    Chloropicrin          1.3           1.3           7.7            <0.1              <0.1             1.4

    Chloral hydrate      79.0          75.0          55.0            <0.1              <0.1            45.0

    TOX                 557.0         540.0         339.0            59.0              27.0           379.0
                                                                                                                  

    a  Clark et al. (1994). 
    b  TOC = 3.0 mg/litre; pH = 7.6.

    Table 10. Summary of impact of water quality and treatment variables on disinfectant by-product formation

                                                                                                                                              
    Variable       Impact on THMs          Impact on HAAs          Impact on aldehydes        Impact on                Impact on bromate
                                                                                              chlorate/chlorite
                                                                                                                                              

    Contact time   Curvilinear increase    Curvilinear increase    Linear increase as long    Linear increase in       Curvilinear increase 
                   with increasing         with increasing         as residual chemical       bleach solutions         with most bromate 
                   contact time            contact time            present                    No discernible effects   forming in <5 min
                   Rapid formation <5 h    Rapid formation <5 h    Secondary reactions        in dilute solutions      Formation is a function 
                   90% formation in 24 h   90% formation in 24 h   between disinfectants      If oxidation of          of ozone residual and 
                   Levels off at 96 h      Levels off at 150 h     and aldehydes possible     hypochlorite, contact    bromide
                                                                                              time has a positive 
                                                                                              effect

    Disinfectant   Rapid and curvilinear   Curvilinear increase    Curvilinear with           Concentrations related   Linear increase after 
    dose           increase after TOC      after TOC demand with   increasing ozone dose      to hypochlorite doses    TOC demand and then 
                   demand with dose,       increasing dose,        or chlorine dose           applied                  levelling off after 
                   levelling off at 2.0    levelling off at 2.0    No appreciable effect      Ozone oxidation of       ozone residual 
                   mg/litre for TOC of     mg/litre                after ozone/DOC = 2 : 1    hypochlorite increases   disappearance
                   2.0 mg/litre                                                               with dose

    pH             Curvilinear increase    Mixed, possible pH      Negative effect (forms     Positive effect          Strong linear positive 
                   with increasing pH to   maximum for DCAA        mostly through molecular   Decomposition of         effect
                   pH 7.0 and possible     and DBAA                ozone)                     hypochlorite increases   Hydroxyl radical 
                   pH maximum              TCAA decreases up       25% decrease for pH        with pH                  generation efficiency 
                   No positive effect at   to pH > 9               7-8.5                      Oxidation of             increases
                   pH > 9.5                DCAA maximum at pH                                 hypochlorite by ozone 
                                           7-7.5                                              increases

    Temperature    Linear increase with    Linear increase with    Terminal products such     Positive effect          Curvilinear increase 
                   increasing temperature  increasing temperature  as carbon dioxide          Decomposition of         20-30% increase for 
                   (10-30 C; 15-25%       (10-30 C; 20-30%       increase and total         hypochlorite increases   15-25 C
                   increase)               increase)               aldehydes slightly 
                                                                   decrease

                                                                                                                                              

    Table 10. (continued)

                                                                                                                                              
    Variable       Impact on THMs          Impact on HAAs          Impact on aldehydes        Impact on                Impact on bromate
                                                                                              chlorate/chlorite
                                                                                                                                              

    TOC            Increase with           Increase with           Positive effect            Negative effect if       Decreases with 
                   increasing TOC;         increasing TOC;         (hydrophobic fraction      ozone is used for        increasing TOC; 
                   precursor content       precursor content       mostly responsible)        hypochlorite oxidation   precursor content 
                   important               important               Doubles for every          Most likely no effect    important
                   Humic acids more        Humic acids more        2 mg/litre                 with hypochlorite        Non-humic acid being 
                   reactive than fulvic    reactive than fulvic                                                        less reactive 
                   acids                   acids                                                                       with ozone

    UVA254         Increase with           Increase with           Positive effect            Negative effect if       Decreases with 
                   increasing UV           increasing UV           Ozone demand               ozone is used for        increasing UV 
                   absorbance; precursor   absorbance;             increases with UV          hypochlorite oxidation   absorbance; precursor 
                   content important       precursor content       (UV absorbance is          Probable negative        content important
                   Aromaticity of TOC      important               mostly due to              effect with              Humic acid being 
                   being more important    Aromaticity of TOC      aromaticity and            hypochlorous acid        more reactive with 
                                           being more important    hydrophobic fraction)                               ozone

    Bromide        Shift towards           Shift towards           Independent of bromide     Shift towards more       Bromide threshold 
                   brominated species      brominated species      at <0.25 mg/litre          toxic bromate in         Curvilinear increase 
                                                                   At >0.25 mg/litre,         hypochlorite solutions   and dependent upon 
                                                                   aldehydes can decrease                              ozone residual
                                                                   due to ozone-bromide 
                                                                   oxidation

    Alkalinity     No discernible effect   No discernible effect   Slight positive effect     Unknown                  Positive effect

    Minimization   TOC removal,            TOC removal,            pH control, TOC            Avoid hypochlorite       pH depression, 
    strategies     minimizing chlorine     minimizing chlorine     removal by coagulation,    dosing solution          ammonia addition, 
                   residual, alternative   residual, alternative   GAC, optimizing doses,     Minimize storage         radical scavengers, 
                   disinfectants, pH       disinfectants, pH       contact time               Properly tune            minimizing and 
                   control, minimizing     control, minimizing                                generators               optimizing ozone 
                   contact time            contact time                                       Use freshly made         residual
                                                                                              solutions
                                                                                                                                              

    Table 10. (continued)

                                                                                                                                              
    Variable       Impact on THMs          Impact on HAAs          Impact on aldehydes        Impact on                Impact on bromate
                                                                                              chlorate/chlorite
                                                                                                                                              

    Removal        GAC, electron beam,     GAC, electron beam      Biofiltration, advanced    Ferrous sulfate, GAC,    Ferrous sulfate, UV 
    strategies     air stripping                                   oxidation, GAC,            electron beam, UV        irradiation, 
                                                                   nanofilters                irradiation,             high-energy electron 
                                                                                              nanofilters              beam, GAC
                                                                                                                                              
    

    oxidation, membrane processes, carbon adsorption and biological
    degradation. For many organic compounds that are difficult to oxidize,
    such as chloroform, the kinetics of ozone oxidation are generally very
    slow but are faster if used in combination with UV irradiation. GAC
    adsorption and membrane filtration are relatively expensive processes;
    moreover, NOM removal by GAC cannot be accomplished to any significant
    degree in a filter/adsorber (i.e., GAC filter cap) mode but requires a
    separate post-filtration adsorber bed. The use of membranes requires
    pretreatment to prevent fouling, as well as processing of waste brine.
    The use of ozone in combination with biologically active GAC filters
    is a promising alternative to reduce DBP precursors.

    2.11.3  Inorganic by-products

         Properly designed and operated chlorine and chlorine dioxide
    generator systems can minimize some of the production of chlorate ion.
    Removal of chlorite and chlorate has been reported using reduction by
    Fe2+ or sulfite or by GAC (Voudrias et al., 1983; Lykins & Clark,
    1988). GAC is seen as problematic because of chlorate production and a
    short bed life. A chemical process using an appropriate agent such as
    reduced iron (e.g., ferrous sulfate) appears to be a more promising
    approach (Kraft & van Eldick, 1989; Gordon et al., 1990).

         If bromate is present in treated water entering the coagulation
    process (i.e., formed during pre-ozonation), several options exist for
    its removal. An aqueous-phase reducing agent (e.g., Fe2+) can be
    added at the rapid mix step. Powdered activated carbon can likewise be
    added as a solid-phase reductant to remove bromate and DBP precursors.
    Not all utilities contemplating ozone application intend to employ
    pre-ozonation. Rather, they may use intermediate ozonation prior to
    the filtration process; in this situation, removal of bromate by
    activated carbon is possible. This approach has potential relevance to
    integration of GAC columns into a process train or, more
    realistically, to retrofitting of rapid sand filters with GAC filters.
    For groundwaters that require no coagulation, bromate can be removed
    after ozonation using a GAC filter, UV irradiation or high-energy
    electron beam irradiation (Siddiqui et al., 1994, 1996a,b,c).

         Brominated or bromochlorinated amines formed during the oxidation
    step of the process train using chlorine can potentially be removed
    using a suitable activated carbon before terminal chlorination.
    However, carbon that has an accumulation of surface oxides, which
    develop through reaction of amines, will have a diminished capacity to
    reduce halogenated amines to nitrogen. Organic amines can potentially
    be removed by activated carbon adsorption.

    2.11.4  Organic by-products

         There are some technologies for removing organic contaminants
    formed after chlorination and chloramination, a less viable option
    than minimizing their formation through DBP precursor removal or use
    of alternative disinfectants. Studies of ozone oxidation have shown
    that aromatic compounds, alkenes and certain pesticides (some of which
    have structural similarities to certain organic DBPs) are removed well

    by ozone treatment, but that alkanes are poorly removed. Also, removal
    efficiency improves for the alkenes and aromatic compounds with
    increasing ozone dosage and for some alkanes with increasing pH. For
    most compounds, the efficacy of ozone is not affected by the
    background water matrix if the ozone is used after coagulation.
    Andrews et al. (1996) showed that using hydrogen peroxide at 1.0
    mg/litre in combination with UV effectively reduced or prevented the
    formation of aldehydes.

    2.12  Models for predicting disinfectant by-product formation

         The regulation of THMs and other halogenated DBPs has been
    complicated by findings that alternative disinfectants to free
    chlorine may also form by-products that are of potential health
    concern. Additional complicating factors impacting the regulation of
    DBPs have been the emergence of  Giardia and  Cryptosporidium as
    major waterborne pathogens.

         In view of the finding that water chlorination produces DBPs,
    some of which are carcinogenic, mutagenic or possibly teratogenic,
    several countries have recently laid down standards for various DBP
    levels. This stimulated the search for mathematical models to describe
    or predict DBP formation in disinfected water and to evaluate the
    effectiveness of water treatment technologies designed to reduce DBP
    levels so as to comply with the standards. Most of these models are
    based on fitting mathematical equations to various empirical
    observations, rather than mechanistic and kinetic considerations. This
    is mainly due to the complexity of the reactions between organic
    precursors and disinfecting chemicals, which usually involve several
    parallel pathways leading to a great variety of products. The
    complexity of the DBP formation reactions also makes it difficult to
    develop universally applicable models for simulating DBP formation
    potential associated with disinfection of a diverse array of natural
    source waters. However, the analysis presented by many models suggests
    that many waters exhibit comparable general responses to changes in a
    given parameter (i.e., responses lending themselves to simulation by a
    particular mathematical functionality), although specific responses
    associated with individual waters may vary. The multiple regression
    models developed by many researchers represent a rational framework
    for modelling DBP formation in many sources. Another potential
    application is the modelling of DBP mixtures, e.g., predicting HAA
    levels from THM and water quality data.

    2.12.1  Factors affecting disinfectant by-product formation and
            variables of interest in disinfectant by-product modelling

         The information on the factors controlling DBP formation, which
    is available in the literature, is briefly summarized below.

         The extent of formation of DBPs is dependent on several water
    quality parameters, such as TOC concentration, UVA254, bromide
    concentration and temperature. It is also dependent on chlorination
    conditions, such as chlorine dose, pH, ammonia concentration and
    contact time. After the various statistically significant factors were

    identified, mathematical equations were developed to describe the
    formation of various DBPs. A least squares method was used to
    determine the optimum equation coefficients that best describe the
    experimental data. The optimum coefficients have been defined as those
    that produce a minimum residual error between the mathematical
    predictions and the experimental data.

    2.12.2  Empirical models for disinfectant by-product formation

         Numerous models for predicting THM formation through chlorination
    have been reported (Moore et al., 1979a; Kavanaugh et al., 1980;
    Engerholm & Amy, 1983; Urano et al., 1983; Amy et al., 1987, 1998;
    Morrow & Minear, 1987; AWWARF, 1991; Hutton & Chung, 1992). Of these,
    models reported by AWWARF (1991) and Amy et al. (1998) are more recent
    and were derived from a variety of natural source waters and more
    realistic treatment conditions. Not much information has been reported
    on the formation of other chlorination DBPs. Only Amy et al. (1998)
    summarized empirical models for THMs, HAAs and chloral hydrate. These
    chlorination by-product models can be used to assess both in-plant and
    distribution system formation of THMs, HAAs and chloral hydrate. Water
    quality conditions such as DOC, pH, temperature and bromide are needed
    as inputs to the models; such data then allow assessment of
    chlorination DBP formation as a function of reaction time:

         DBP concentration (total THMs or THM species, total HAAs or HAA
         species, or chloral hydrate) =
         f(TOC, bromide, chlorine, pH, temperature, time)

         Relatively little is known about the kinetics of the formation of
    bromate and other DBPs during ozonation and the quantitative effects
    of water quality factors (temperature, pH, etc.); such an
    understanding is crucial for evaluating various bromate control
    strategies. Siddiqui & Amy (1993) and Amy et al. (1998) developed
    statistical relations to predict the concentrations of various ozone
    DBPs, including bromate, as a function of water treatment variables.
    Correlation matrix analysis has shown that ozone dose, dissolved ozone
    concentration, bromide concentration, pH and reaction time all have a
    positive influence on bromate formation. Von Gunten & Hoigne (1994)
    developed kinetic models for bromate formation.

         Ozone, as a result of its strong oxidizing power, produces a
    variety of organic by-products, such as aldehydes and ketoacids, when
    used to treat natural source waters. These by-products -- especially
    aldehydes -- are highly biodegradable, and there is concern for
    regrowth of microorganisms following ozone treatment. They are also
    potentially hazardous and may produce increased amounts of chlorinated
    by-products upon chlorination. Siddiqui et al. (1997) developed a
    model to estimate the potential for total aldehyde formation in source
    waters upon ozonation.

    2.12.3  Models for predicting disinfectant by-product precursor
            removal

         It is recognized that chlorination will continue to be the most
    common disinfection process; hence, enhanced removal of DBP precursors
    present in raw sources represents a valuable option for reducing the
    potential for by-product formation. The removal of NOM can be achieved
    either by providing additional processes, such as GAC and
    nanofiltration, or by enhancing the existing coagulation, flocculation
    and sedimentation processes. Predictive models have been developed for
    assessing coagulation efficiency in removing NOM and reducing DBP
    precursor levels (AWWARF, 1991; Amy et al., 1998).

         Coagulation can reduce DOC and DBP precursors but not bromide
    levels; hence, a greater proportion of brominated DBP species can
    potentially be produced in the finished water.

         The effects of precursor removal by chemical coagulation can be
    assessed through the use of treated water models. One can either
    predict DBPs formed under a given degree of precursor removal or
    define the degree of precursor removal required to meet DBP
    regulations. The impact of bromide ion on meeting regulations can also
    be assessed. If one makes the assumption that precursor reactivity
    (i.e., DBP/DOC) does not change, one can also assess other precursor
    removal processes, such as GAC or membrane processes, through use of
    the raw/untreated water models. Care should be exercised when using
    models to approximate post-chlorination DBPs following an ozonation
    step.

    2.13  Summary

    *  The primary and most important role of drinking-water treatment is
       to remove or inactivate harmful microorganisms. Another role is to
       minimize the concentrations of disinfectants and DBPs without
       compromising in any way the removal or inactivation of pathogens.

    *  Drinking-water utility managers must be more knowledgeable about
       options to meet regulations. It is often more practical to use
       treatment methods that control the concentration of several
       contaminants than to modify treatment practices for each new
       standard that is promulgated. 

    *  A thorough understanding of DBP formation would help the successful
       balancing of appropriate microbial inactivation with the
       minimization of DBPs. Water quality variables affect DBP formation
       and must be considered when developing a strategy to control DBPs
       with various disinfectants.

    *  The chemistry of chlorine and its by-products has been well
       studied, and ozone and its by-products have recently received much
       attention. Studies of chlorine dioxide and chloramines and their
       by-products are relatively few, although more work in these areas
       is now being undertaken.

    *  One of the simplest processes to minimize halogenated DBP formation
       is limiting the free chlorine contact time by using monochloramine
       to maintain a distribution system residual following primary
       disinfection by chlorine or ozone. Chloramines are an effective
       means of controlling DBPs. However, the growth of nitrifying
       bacteria (and related production of nitrite) is a potential problem
       in chloraminated water supplies.

    *  Various nitrogen-containing organic compounds may be present in
       source waters after chlorination and chloramination. Because of
       analytical complexities, very few detailed studies have been
       undertaken to determine the individual compounds present and their
       concentrations.

    *  Many factors between the source and the tap can influence the DBPs
       to which consumers are exposed. Although THMs and HAAs continue to
       form with increasing contact time, some other halogenated DBPs,
       such as HANs and haloketones, form rapidly but then decay in the
       distribution system as a result of hydrolysis. This has major
       implications regarding exposure to these DBPs, depending upon their
       proximity to the treatment plant. For treated source waters, median
       levels of HAAs are often approximately one-half of the median THM
       levels.

    *  For low-bromide source waters, chloroform is normally the dominant
       THM species; DCA and TCA are the most prevalent HAA species; DCAN
       is the most prevalent HAN species; and 1,1,1-TCPN is the most
       prevalent of the two measured haloketones. Very low levels of
       chloropicrin have been observed by various researchers; this
       compound appears to form slowly during the incubation period, with
       concentrations tending to level off at 40 h. For high-bromide
       waters, increased levels of brominated DBPs are observed.

    *  Chlorine dioxide is a strong oxidant that under certain conditions
       surpasses chlorine in its ability to destroy pathogenic organisms.
       When chlorine dioxide is prepared and administered without excess
       free chlorine, THMs and other chlorinated by-products are not
       produced, but inorganic by-products are formed. 

    *  TOC levels have been found to be correlated with halogenated DBP
       formation. The nature of this relationship varies with the source.
       TOC removal can be used as a surrogate for the reduction of DBP
       formation.

    *  Although the presence of chloral hydrate and HANs in chlorinated
       samples may be attributed to precursors other than amino acids, the
       potential for amino acids to be present in natural sources is well
       documented. Surface waters, but not groundwaters, tend to contain
       amino acids. However, the removal of these precursors during
       conventional water treatment is not well understood.

    *  The amount of chlorate that is present in delivered hypochlorite
       solutions depends on many factors. Freshly made hypochlorite
       solutions will contain less chlorate than hypochlorite that is
       stored without concern for temperature and pH. If a utility is
       using a single tank to store hypochlorite, it is likely that the
       level of chlorate is increasing in the tank. Thus, storage tanks
       should be periodically flushed and cleaned, and, if possible, the
       storage time should be reduced.

    *  Models have been developed that can be used to simulate the fate
       and movement of DBP precursors in distribution systems. The models
       can be designed as a planning tool for evaluating the impacts of
       source water management strategies and estimating DBP exposures.
       Some limitations of existing models include calibration with a
       limited database, application to only a specific water source or
       group of related sources, lack of terms to simulate important
       parameters, such as reaction time, and inadequate validation.

    3.  TOXICOLOGY OF DISINFECTANTS

         In assessing the hazards associated with drinking-water
    disinfection, it is important not to neglect the disinfectants
    themselves. Adding disinfectant in excess of the demand has several
    practical benefits. First, it ensures that reaction of the
    disinfectant with DBP precursors (largely organic material and
    ammonia) does not shorten contact time to the point of ineffective
    disinfection. Second, residual disinfectant helps to prevent regrowth
    of organisms in the remaining portions of the treatment and
    distribution systems.

         The result of this practice, however, is that one of the
    chemicals that is present in the finished water at the highest
    concentration is the disinfectant. In the present regulatory climate
    in many countries, chemicals that are introduced as direct additives
    to food would be subjected to a significant amount of toxicological
    screening before they could be used. Since the major disinfectants
    were introduced almost 100 years ago, they were subjected to much less
    thorough toxicological evaluations than would be required today.
    However, many of these data gaps have been addressed in the past
    decade.

    3.1  Chlorine and hypochlorite

    3.1.1  General toxicological properties and information on 
           dose-response in animals

         Chlorine gas has long been recognized as a lung irritant. This
    topic will not be reviewed in the present document, as it appears to
    be largely irrelevant to the small amounts of chlorine that are
    volatilized from chlorinated water in showers or other points of use
    in the household. In water treatment plants, however, there is a
    possibility of occupational exposures that could have severe sequelae.
    For information on these higher-level exposures, the interested reader
    is referred to a recent review by Das & Blanc (1993). The effects of
    chlorine gas that have been observed in humans will be discussed in
    section 3.1.3.

         Sodium hypochlorite (NaOCl) or calcium hypochlorite (Ca(OCl)2)
    solutions have also been utilized extensively in the disinfection of
    drinking-water. The stock solutions used for this purpose are highly
    caustic and are a clear concern for occupational exposures. The
    concentration required to produce irritation and decreased basal cell
    viability in the skin of guinea-pigs after an application period of
    2 weeks was 0.5% sodium hypochlorite (Cotter et al., 1985). Reducing
    the concentration to 0.1% resulted in no effect on basal cell
    viability relative to control animals. Yarington (1970) demonstrated
    that instillation of bleach into the oesophagus of dogs produced
    irritation. The minimal exposure that produced oesophageal burns was
    10 ml of commercial bleach with a 5-min exposure. It should be noted
    that the highly alkaline pH (about pH 11) of sodium hypochlorite is
    not likely to be encountered in drinking-water.

         There have been relatively few evaluations of the effects of
    chlorine or hypochlorite in drinking-water. The present review will
    focus on studies with treatment periods longer than 4 weeks where
    drinking-water was the primary route of exposure. Reference to earlier
    studies of shorter duration and less general applicability to a safety
    evaluation can be found in previous reviews (Bull, 1980, 1982a,b,
    1992; Bull & Kopfler, 1991).

         Daniel et al. (1990a) evaluated the toxicity of solutions of
    chlorine prepared by bubbling chlorine gas into distilled water and
    adjusting the pH to 9.4. The nominal concentrations of chlorine used
    were 0, 25, 100, 175 or 250 mg/litre in distilled water (approximately
    0, 3, 10, 16 or 21 mg/kg of body weight per day). These solutions were
    provided as drinking-water to both male and female Sprague-Dawley rats
    (10 per sex per dose) for 90 days. No deaths occurred in any treatment
    group. However, there were statistically significant decreases in
    drinking-water consumption in females treated with 100 mg/litre and
    higher, probably due to decreased palatability. There were no
    consistent effects of chlorine treatment on organ to body weight
    ratios or clinical chemistry parameters. A no-observed-effect level
    (NOEL) of 10 mg/kg of body weight per day was identified by the
    authors based on reduced body weight gain. However, since this was
    associated with reduced palatability of the drinking-water, it is not
    considered to be a true toxicological end-point.

         The study in rats was followed up with another study in B6C3F1
    mice (Daniel et al., 1991a). Male and female B6C3F1 mice (10 per sex
    per group) were administered 12.5, 25, 50, 100 or 200 mg of chlorine
    per litre of drinking-water for 90 days (calculated mean daily doses
    were 2.7, 5.1, 10.3, 19.8 or 34.4 mg/kg of body weight in males and
    2.8, 5.8, 11.7, 21.2 or 39.2 mg/kg of body weight in females). Spleen
    and liver weights were depressed in males, but not in females, at the
    highest dose rates (100 and 200 mg/litre). There were no other
    consistent indications of target organ effects based on serum enzyme
    concentrations. No gross or microscopic lesions could be related to
    treatment with chlorine.

         Several of the following studies utilized solutions of sodium
    hypochlorite as the treatment chemical. It is now known that such
    solutions can contain very high concentrations of chlorate within a
    short time of their preparation (Bolyard et al., 1993). The extent of
    this contamination has not been reported.

         Hasegawa et al. (1986) examined the effects of much higher
    concentrations (0.025, 0.05, 0.1, 0.2 or 0.4%) of sodium hypochlorite
    (equivalent to 7, 14, 28, 55 and 111 mg/kg of body weight per day)
    administered in drinking-water to male and female F344 rats for
    13 weeks. Twenty rats of each sex were assigned to each experimental
    group. Significant suppression of body weight (as a result of
    decreased consumption of water and food) occurred at 0.2% and above.
    The authors noted slight damage to the liver as indicated by increased
    levels of serum enzymes (not specified) at 0.2% and 0.4% sodium
    hypochlorite in both sexes. No evidence of treatment-related pathology
    was observed in this study or in a 2-year study in which males were

    subjected to 0.05% or 0.1% (13.5 or 27.7 mg/kg of body weight per day)
    and females to 0.1% or 0.2% (34 or 63 mg/kg of body weight per day)
    sodium hypochlorite. The extended exposures were conducted with
    50 animals of each sex per treatment group. Analysis of dosing
    solutions was not reported.

         In a 2-year bioassay, the National Toxicology Program (NTP)
    examined chlorine at 0, 70, 140 or 275 mg/litre (expressed as atomic
    chlorine, Cl) in drinking-water of F344 rats and B6C3F1 mice (70 per
    sex per group) (NTP, 1992). These solutions were prepared from gaseous
    chlorine and neutralized to pH 9 by the addition of sodium hydroxide.
    Stability studies indicated that 85% of the initial target
    concentration remained after 3 days of preparation. Stock solutions
    (concentrations not specified) were prepared once weekly, and
    solutions for drinking were prepared 4 times weekly. Based on body
    weight and water consumption, doses in these studies were
    approximately 0, 4, 7 or 14 mg/kg of body weight per day for male
    rats; 0, 4, 8 or 14 mg/kg of body weight per day for female rats; 0,
    7, 14 or 24 mg/kg of body weight per day for male mice; and 0, 8, 14
    or 24 mg/kg of body weight per day for female mice. The only
    treatment-related non-tumour pathology was found to be a dilatation of
    renal tubules in male mice receiving 275 mg/litre for more than 66
    weeks. No non-neoplastic lesions were observed in either male or
    female rats.

         A number of immunological changes have been associated with the
    treatment of rodents with sodium hypochlorite in drinking-water. Water
    containing 25-30 mg of sodium hypochlorite per litre was found to
    reduce the mean number of peritoneal exudate cells recovered from
    female C57BL/6N mice after 2 weeks of treatment. This was, in turn,
    associated with a significant decrease in macrophage-mediated
    cytotoxicity to melanoma and fibrosarcoma cell lines (Fidler, 1977).
    The treatment period was increased to 4 weeks in a subsequent study,
    which demonstrated that 25 mg of sodium hypochlorite per litre
    decreased the ability of peritoneal macrophages to phagocytose
    51Cr-labelled sheep red blood cells. Macrophages obtained from the
    mice treated with hypochlorite were found to be less effective in
    destroying B16-BL6 melanoma cells  in vitro. Mice so treated were
    also found to have increased pulmonary metastasis of B16-BL6 cells
    when they were introduced by subcutaneous injection (Fidler et al.,
    1982).

         Exon et al. (1987) examined the immunotoxicological effects of
    sodium hypochlorite at 5, 15 or 30 mg/litre (0.7, 2.1 or 4.2 mg/kg of
    body weight per day) in the drinking-water of male Sprague-Dawley rats
    (12 per dose) for 9 weeks. Delayed hypersensitivity reaction to bovine
    serum albumin was observed at the highest dose administered. Oxidative
    metabolism by adherent resident peritoneal cells was decreased at 15
    and 30 mg/litre, and the prostaglandin E2 levels of these cells were
    found to be significantly elevated. No effects on natural killer cell
    cytotoxicity, antibody responses, interleukin 2 production or
    phagocytic activity were observed. The effects on macrophage activity
    suggest that some impairment does occur at relatively low levels of
    sodium hypochlorite. As pointed out by the authors, these were

    relatively mild effects, the significance of which was unknown. It is
    not clear that these effects would be translated into a significant
    impairment of the immune response to a particular infectious agent.
    However, modification of macrophage function appears to be one of the
    most sensitive responses identified in studies of chlorine or
    hypochlorite in experimental animals. A study in which female C57BL/6
    mice were administered hypochlorite in their drinking-water (7.5, 15
    or 30 mg of hypochlorite per litre) for 2 weeks showed no effects on
    the immune system as measured by spleen and thymus weight,
    plaque-forming cell response, haemagglutination titre and lymphocyte
    proliferation (French et al., 1998).

         Altered liver lipid composition has been observed as a result of
    acute intragastric administration of sodium hypochlorite (5 ml of a 1%
    solution) to rats (Chang et al., 1981). These data do not provide a
    clear indication of whether these effects might give rise to
    pathology. The concentrations of hypochlorite utilized were much
    greater than those that would be encountered in drinking-water.

         The effects of hypochlorous acid and hypochlorite on the skin
    have received relatively little attention despite the current interest
    in bathing as a significant source of chemical exposure from
    drinking-water. Robinson et al. (1986) examined the effects of both
    hypochlorite and hypochlorous acid solutions applied to the skin of
    the entire body of female Sencar mice except for the head. Exposures
    were to 1, 100, 300 or 1000 mg/litre as hypochlorous acid at pH 6.5
    for 10 min on 4 consecutive days. Hypochlorite (formed by raising the
    pH to 8.5) was studied only at 1000 mg/litre. Significant increases in
    epidermal thickness and cell counts within the epidermal layer were
    observed at concentrations of hypochlorous acid (pH 6.5) of 300
    mg/litre and above, but the thickness of the skin was not
    significantly different from that in animals at 100 mg/litre. The
    increases in skin thickness were associated with an epidermis whose
    thickness was increased to 4-6 cells as compared with the normal 1-2
    cells seen in mice. The effects of hypochlorite were much less marked.
    Following a single application, the increased thickness of the skin
    observed in mice exposed to hypochlorous acid (i.e., pH 6.5) did not
    appear until 4 days after the treatment. This differed from
    hypochlorite, other disinfectants and the positive control,
    12- O-tetradecanoylphorbol-13-acetate (TPA). In the latter cases, the
    maximal response was observed within 24-48 h after treatment. The
    hyperplastic response to hypochlorous acid required 12 days to return
    to normal. This study suggests a considerable margin of safety between
    the concentrations of chlorine required to produce hyperplasia and
    those that are found in drinking-water.

         The reactive nature of chlorine always raises questions of
    whether it is chlorine or a by-product that is responsible for any
    effect. Several studies have examined the formation of by-products in
    the gastrointestinal tract following the administration of chlorine.
    Invariably, these studies have involved the administration of chlorine
    or hypochlorite by gavage at very high concentrations relative to the
    amounts that would be encountered in chlorinated drinking-water. As a
    consequence, the by-products formed following gavage dosing of high

    concentrations may not be representative of the by-products that would
    be seen following the consumption of modest to moderate levels of
    chlorine in larger volumes of water. A particular issue is that the
    high organic carbon concentration relative to chlorine that would be
    encountered in the gastrointestinal tract when water is consumed at
    low concentrations should dissipate disinfectant before sufficient
    oxidative power would be present to break down substrates to small
    molecules. Despite these design flaws, the data do indicate that
    by-products are formed. The bulk of them remain as higher molecular
    weight products, which may have little toxicological importance.

         Vogt et al. (1979) reported that chloroform could be measured in
    the blood, brain, liver, kidneys and fat of rats to which sodium
    hypochlorite was administered by gavage at doses of 20, 50 or 80 mg in
    5 ml of water. Thus, the by-product chloroform can be formed by the
    reaction of chlorine with stomach contents.

         Mink and co-workers (1983) pursued this observation and found
    that other by-products could be detected in the stomach contents and
    plasma of rats that had been administered sodium hypochlorite
    solutions neutralized to pH 7.9. In addition to chloroform, DCAN, DCA
    and TCA were detected in the stomach contents analysis. DCA and TCA
    were also detected in blood plasma.

         The third group of compounds identified as by-products of
    chlorination in stomach contents of the rat are the organic
     N-chloramines (Scully et al., 1990).  N-Chloroglycine,
     N-chloroleucine or  N-chloroisoleucine and  N-chlorophenylalanine
    were confirmed products of reactions with normal amino acids that
    would ordinarily be found in the gastrointestinal tract.
     N-Chlorovaline and  N-chloroserine were also tentatively
    identified. Organic chloramines are reactive and could be responsible
    for toxic effects that may be attributed to chlorine in toxicological
    studies. The chlorine demand of free amino acids in stomach contents
    was found to be only about 4% of the total. Consequently, this process
    may be substrate-limited at concentrations of chlorine found as
    residuals in drinking-water. However, use of higher concentrations of
    chlorine would also lead to breakdown of proteins present in the
    stomach fluid. Thus, as concentrations are increased to levels that
    would be used in animal studies, these products would be formed at a
    much higher concentration, similar to the phenomena noted with THM and
    HAA by-products.

    3.1.2  Reproductive and developmental toxicity

         In general, animal studies have demonstrated no reproductive or
    teratogenic effects of chlorine. Druckrey (1968) examined the effects
    of water chlorinated to a level of 100 mg/litre (approximately 10
    mg/kg of body weight per day) in BDII rats for seven generations. No
    effects were observed on fertility, growth or survival. 

         A number of subsequent studies have studied the effects of
    chlorine or hypochlorite on more specific aspects of reproduction or
    development. Meier et al. (1985b) reported that oral administration of

    sodium hypochlorite (pH 8.5) prepared from chlorine gas and
    administered at 4 or 8 mg/kg of body weight per day for 5 weeks
    increased the incidence of sperm head abnormalities in B6C3F1 mice
    (10 animals per group). The effect was not observed when the solutions
    were administered at pHs at which hypochlorous acid was the
    predominant species (pH 6.5). However, other studies have not been
    able to associate adverse reproductive outcomes with the
    administration of chlorine or sodium hypochlorite.

         Carlton et al. (1986) found no evidence of sperm head
    abnormalities or adverse reproductive outcomes in Long-Evans rats.
    Male rats were treated for 56 days prior to mating and female rats
    from 14 days prior to mating through gestation. Each experimental
    group consisted of 11-12 males and 23-24 females. Solutions of
    chlorine were prepared at pH 8.5, so the study evaluated hypochlorite
    as the dominant form in the drinking-water. Doses were as high as
    5 mg/kg of body weight per day. 

    3.1.3  Toxicity in humans

         There have been significant human exposures to chlorine and
    hypochlorite solutions. Much of that experience is with inhalation of
    chlorine gas, which is known to be a strong respiratory irritant.
    Chlorine gas is also the largest single component involved in toxic
    release incidents. A third major source of exposure is solutions of
    sodium hypochlorite, usually marketed as bleach. Bleach is frequently
    involved in human poisonings. These exposures are not particularly
    relevant to exposures to chlorine or hypochlorite in drinking-water.
    Therefore, only a few case reports are identified that illustrate the
    types of problems that have been encountered. There was no attempt to
    make this review comprehensive.

         The irritating effects of chlorine gas have been well documented
    because of its use as a chemical warfare agent during World War I (Das
    & Blanc, 1993). In a follow-up of survivors of gassing, it was
    concluded that there was no evidence of permanent lung damage;
    however, these studies clearly indicated that survivors had breath
    sounds that suggested bronchitis and limited chest and diaphragmatic
    movement, even emphysema. Most studies suggested that there were high
    incidences of acute respiratory disease and a lesser prevalence of
    chronic sequelae. Similar sequelae have been identified following
    exposure of humans to accidental releases of chlorine gas. In these
    more modern characterizations, the acute signs and symptoms included a
    high incidence of pulmonary oedema and severe bronchitis. These signs
    and symptoms are of generally short duration and resolve themselves
    over the course of about 1-4 weeks. However, chronic sequelae are
    observed in some individuals, depending in part upon the severity of
    the exposure. In such cases, a decrease in the forced expiratory
    volume is the most consistently reported clinical sign. 

         Two recent reports suggest that chronic sequelae to acute
    exposures to chlorine gas may be more prevalent than previously
    appreciated. Moore & Sherman (1991) reported on an individual who was
    previously asymptomatic and who developed chronic, recurrent asthma

    after exposure to chlorine gas in an enclosed place. Schwartz et al.
    (1990) followed 20 individuals who had been exposed to chlorine gas in
    a 1975 incident. The prevalence of low residual lung volume was
    increased during the follow-up period. Sixty-seven per cent of those
    tested were found to have residual volumes below 80% of their
    predicted values. Five of 13 subjects tested for airway reactivity to
    methacholine were found to have a greater than 15% decline in forced
    expiratory volume.

         Controlled studies have been conducted in healthy, non-smoking
    men exposed to chlorine gas at 1.5 and 2.9 mg/m3 (0.5 and 1.0 ppm)
    for 4 or 8 h (reviewed in Das & Blanc, 1993). Four hours of exposure
    to 2.9 mg/m3 (1.0 ppm) produced significant decreases in the forced
    expiratory volume. One individual who was found to be experiencing
    more difficulty than other subjects at this dose and who was later
    exposed to 1.5 mg/m3 (0.5 ppm) experienced a significant decrease in
    forced expiratory volume. While 1.5 mg/m3 (0.5 ppm) appears
    protective for most people, some more sensitive individuals may in
    fact have more significant responses to chlorine gas. 

         The effects of chronic exposure to chlorine gas have received
    only limited study. In one study of paper mill workers, a more rapid
    age-related decrease in lung volumes of workers exposed to chlorine
    relative to those exposed to sulfur dioxide was noted, but the trend
    was not statistically significant (Das & Blanc, 1993). Other studies
    failed to identify chronic sequelae.

         There are frequent reports of human poisonings from bleach. Most
    often these exposures result from the mixing of bleach with acidic
    products or ammonia. Acidification converts hypochlorite to
    hypochlorous acid, which dissociates to chlorine gas, offgasses very
    rapidly from the solutions and presents an inhalation exposure (MMWR,
    1991). Mixing bleach with ammonia results in the formation of
    monochloramine and dichloramine, both of which are effective
    respiratory irritants (MMWR, 1991).

         Any potential effects of chlorine or hypochlorite in
    drinking-water are obscured by the fact that by-products inevitably
    coexist with the residual chlorine. One series of studies in which
    by-products formed were minimized by dissolving chlorine in distilled
    water attempted to identify effects of chlorine in drinking-water on
    humans. Chlorine in drinking-water was administered in a rising-dose
    tolerance study beginning with 0.1 mg/litre in two 500-ml portions and
    rising to a concentration of 24 mg/litre, equivalent to 0.34 mg/kg of
    body weight per day (Lubbers & Bianchine, 1984). No clinically
    important changes were observed. No findings of clinical importance
    were identified in a follow-up treatment with repeated dosing with
    500-ml portions of a solution containing 5 mg of chlorine per litre
    for a 12-week period (Lubbers et al., 1984a).

         Another study attempted to determine whether consumption of
    chlorinated drinking-water affected blood cholesterol levels (Wones et
    al., 1993a). The impetus for this study was a toxicological study in
    pigeons that suggested that chlorine raised blood cholesterol levels

    and modified serum thyroid levels (Revis et al., 1986a,b) and an
    epidemiological study that associated small increases in cholesterol
    of women with residence in communities having chlorinated water
    (Zeighami et al., 1990a,b; described in detail in section 5.2.2). A
    prior study (quoted in Wones et al., 1993a) was conducted that
    examined men who consumed water containing 2, 5 or 10 mg of chlorine
    per litre and found a small increase in serum cholesterol levels at
    the highest dose group. However, no control group was studied, so the
    changes could have been attributed to the change in diet imposed as
    part of the study protocol (Wones & Glueck, 1986). The longer-term
    study was composed of 30 men and 30 women who received a controlled
    diet for the duration of the study. The first 4 weeks represented an
    acclimatization period during which all subjects received distilled
    water. Half the subjects were assigned to a group that consumed 1.5
    litres of water containing 20 mg of chlorine per litre for the
    following 4 weeks. At the end of each 4-week period, blood was
    analysed for cholesterol, triglycerides, high-density lipoprotein
    (HDL) cholesterol, low-density lipoprotein (LDL) cholesterol or
    apolipoproteins A1, A2 and B. There were no significant effects. There
    was a slight trend towards lower thyroid hormone levels in men
    consuming chlorine, but this was not clinically significant (Wones et
    al., 1993a). These data suggest that observations obtained previously
    in pigeons could not be repeated under comparable conditions of
    chlorine consumption. It is notable that the animals utilized in the
    original pigeon study had consumed a modified diet (Revis et al.,
    1986a) that was deficient in calcium and other trace metals. A
    subsequent study failed to replicate the previous results in pigeons
    (Penn et al., 1990).

    3.1.4  Carcinogenicity and mutagenicity

         The International Agency for Research on Cancer (IARC) has
    evaluated the carcinogenicity of hypochlorite salts and concluded that
    there were no data available from studies in humans on their
    carcinogenicity and inadequate evidence for their carcinogenicity in
    experimental animals. Hypochlorite salts were assigned to Group 3: the
    compounds are not classifiable as to their carcinogenicity to humans
    (IARC, 1991).

         Several studies have shown that sodium hypochlorite produces
    mutagenic responses in bacterial systems and mammalian cells
     in vitro. However, there is no evidence of activity in mammalian
    test systems  in vivo. It is not clear to what extent this is
    influenced by the formation of mutagenic by-products as a result of
    reactions with components of the incubation media. Wlodkowski &
    Rosenkranz (1975) used short-term exposures of  Salmonella 
     typhimurium strain TA1530 followed by ascorbic acid-induced
    decomposition to reduce the cytotoxic effects of hypochlorite. The
    investigators applied 0.14 mol per tube and added ascorbic acid after
    intervals of 5, 10 and 15 min. At 5 min, a clear positive response was
    observed with minimal cytotoxicity. Significant responses were also
    observed in strain TA1535, but not in strain TA1538.

         Rosenkranz (1973) and Rosenkranz et al. (1976) also demonstrated
    a positive mutagenic response in DNA polymerase A deficient
     Escherichia coli to 0.006 mol of sodium hypochlorite. This response
    was unaffected by the addition of catalase, suggesting that the
    response was not related to the generation of hydrogen peroxide.

         Matsuoka et al. (1979) reported that sodium hypochlorite at a
    concentration of 6.7 mmol/litre (0.5 mg/ml) produced chromosomal
    aberrations in Chinese hamster ovary (CHO) cells in the presence of S9
    mix. This concentration was cytotoxic in the absence of S9. Some
    concern must be expressed about whether responses observed with such
    high and clearly cytotoxic concentrations in an  in vitro system
    represent specific clastogenic effects. The authors report only one
    concentration tested with and without S9. It is probable that the
    positive response in the presence of S9, if it is a specific response,
    was a result of detoxifying hypochlorite. This protection could be
    non-specific as well, in that it may not have depended upon any
    catalytic activities present in the S9 fraction (i.e., the added
    protein may have acted as a reactive sink to dissipate excess
    hypochlorite). Consequently, it is difficult to use these data in
    interpreting the effects of chlorine or hypochlorite  in vivo.

         Ishidate (1987) studied the induction of chromosomal aberrations
    in cultures of Chinese hamster CHL cells at sodium hypochlorite
    concentrations ranging from 125 to 500 g/ml without exogenous
    metabolic activation and from 31 to 125 g/ml with and without rat
    liver S9 mix. A clear increase in the number of cells with structural
    chromosomal aberrations was observed at 500 g/ml without S9 mix,
    while the results obtained in the other series, showing weakly
    positive responses, were considered inconclusive.

         Meier et al. (1985b) evaluated the ability of hypochlorite and
    hypochlorous acid to induce chromosomal damage or micronuclei in the
    bone marrow of CD-1 mice. The samples to be tested were generated by
    bubbling chlorine gas into water and then adjusting the pH to 6.5
    (predominantly hypochlorous acid) or 8.5 (predominantly hypochlorite).
    The doses administered were 1.6, 4 or 8 mg/kg of body weight for 5
    consecutive days. There was no evidence of increased micronuclei or
    chromosomal abnormalities in bone marrow cells. Significant positive
    responses were observed with positive control chemicals in both
    assays. As reported in section 3.1.2, these authors detected a
    positive response in the sperm head abnormality assay in mice treated
    at these same doses of hypochlorite in two separate experiments. This
    assay is used primarily as a mutagenicity assay rather than as an
    assay for reproductive toxicities. Hypochlorous acid had no effect in
    the sperm head abnormality assay.

         Tests of the ability of hypochlorite to induce cancer in rodents
    were conducted in F344 rats by Hasegawa et al. (1986). Sodium
    hypochlorite concentrations of 0, 500 or 1000 mg/litre (males) and 0,
    1000 or 2000 mg/litre (females) were administered in the
    drinking-water for 104 weeks (equivalent to 13.5 and 27.7 mg/kg of
    body weight per day for males and 34.3 and 63.2 mg/kg of body weight
    per day for females). There were 50 male and 50 female rats assigned

    to each experimental group. No tumours could be attributed to sodium
    hypochlorite administration.

         NTP (1992) conducted a 2-year bioassay of chlorine in F344 rats
    and B6C3F1 mice. The concentrations administered in drinking-water
    were 0, 70, 140 or 275 mg/litre, and there were 70 animals of each sex
    assigned to each group (approximately 0, 4, 8 or 14 mg/kg of body
    weight per day for rats and 0, 7, 14 or 24 mg/kg of body weight per
    day for mice). There was an apparent positive trend in the induction
    of stromal polyps of the uterus of female mice treated with chlorine,
    but this was considered unlikely to be treatment-related because the
    incidence was below those observed in historical controls. In female
    rats, there was an increase in mononuclear cell leukaemia at both 140
    and 275 mg/litre (8 and 14 mg/kg of body weight per day). However, the
    response was not considered treatment-related because it fell within
    the range of historical controls, there was no apparent dose-response,
    and there was no evidence for such an increase in male F344 rats.

         A single study suggested that sodium hypochlorite could act as a
    promoter of skin tumours following initiation with
    4-nitroquinoline-1-oxide in female ddN mice (Hayatsu et al., 1971). A
    solution of sodium hypochlorite that contained 10% effective
    concentrations of chlorine was utilized. Skin tumours were produced in
    9 of 32 mice given 45 applications of sodium hypochlorite following
    initiation. Sodium hydroxide solutions were utilized as a control for
    the alkaline pH of sodium hypochlorite and produced no tumours. No
    tumours were observed with 60 applications of sodium hypochlorite
    solution in non-initiated mice. Pfeiffer (1978) conducted a much
    larger experiment that utilized 100 mice per group. This author found
    that a 1% sodium hypochlorite solution applied alternately with
    benzo [a]pyrene for 128 weeks was ineffective in producing skin
    tumours in female NMRI mice above those that had been initiated with
    benzo [a]pyrene alone at doses of 750 or 1500 g. Pretreatment with
    the sodium hypochlorite solution before application of the
    benzo [a]pyrene actually reduced tumour yields at 128 weeks with
    doses of either 750 or 1500 g of benzo [a]pyrene. Sodium
    hypochlorite used in a more traditional initiation/promotion study
    (i.e., sodium hypochlorite treatment following initiation with
    benzo [a]pyrene) produced a decrease in the tumour yield with the 750
    g dose of benzo [a]pyrene, but had no effect following 1500 g.
    Thus, the ability of sodium hypochlorite to act as a tumour promoter
    may depend upon the initiator used, or the smaller experiment of
    Hayatsu et al. (1971) may simply be a false result.

         As pointed out in section 3.1.1, application of solutions of
    hypochlorous acid to the skin of Sencar mice results in the
    development of hyperplasia. The concentrations required are
    considerably lower (300 mg/litre) (Robinson et al., 1986) than those
    used in the studies of either Hayatsu et al. (1971) or Pfeiffer
    (1978). Sodium hypochlorite was also effective at lower doses, but
    less so than equivalent concentrations of hypochlorous acid. These
    results suggest that these prior evaluations may have been conducted
    at too high a dose. There appear to be no reports on the effectiveness

    of hypochlorous acid as a tumour promoter, but the lack of activity at
    doses of less than 300 mg/litre would suggest that this is of no
    concern.

    3.1.5  Comparative pharmacokinetics and metabolism

         A series of pharmacokinetic studies using 36Cl-labelled
    hypochlorous acid were conducted by Abdel-Rahman and co-workers
    (1983). These studies are of limited value because the form of 36Cl
    could not be determined in various body compartments. 

    3.1.6  Mode of action

         There are no specific toxicities of chlorine for which a
    mechanism needs to be proposed. It is a strong oxidizing agent, and it
    must be presumed that damage induced at high doses by either gaseous
    chlorine or solutions of hypochlorite is at least partially related to
    this property. In studies in which sodium hypochlorite is used without
    neutralization, a strong alkaline pH can also contribute to its
    effects. There is always the possibility that chlorine is inducing
    subtle effects by virtue of its reaction with organic compounds that
    are found in the stomach. Such reactions have been demonstrated, but
    there is no convincing evidence to date that any specific toxicity can
    be attributed to these by-products.

    3.2  Chloramine

    3.2.1  General toxicological properties and information on 
           dose-response in animals

         There have been relatively few evaluations of the toxic
    properties of chloramine in experimental animals. In large part this
    is because it is not marketed as a product but is created for
    disinfection purposes on-site and  in situ. Chloramine is primarily
    used as a residual disinfectant in the distribution system. The final
    solution consists of mostly monochloramine, with traces of other
    chloramines, such as dichloramine. Chloramines, as a group, are
    generally recognized as potent respiratory irritants, because the
    formation of these compounds when household bleach and ammonia are
    mixed results in a number of poisoning cases each year (MMWR, 1991).
    In spite of this, there has been no attempt to quantify dose-response
    relationships in animals.

         Eaton et al. (1973) investigated concerns about
    chloramine-induced methaemoglobin formation in kidney patients
    dialysed with chloramine-containing water. This was done by examining
    the ability of relatively large volumes of tapwater to oxidize
    haemoglobin in dilute suspensions of red blood cells. This
    circumstance is reflective of dialysis, but not of the concentrated
    suspension of these cells  in vivo. Nevertheless, the authors were
    able to show that methaemoglobin formation occurred in a dose-related
    manner when 1 volume of red blood cells (human) was suspended in 100
    volumes of tapwater containing 1 mg of chloramine per litre or above.
    This effect was not produced by comparable concentrations of sodium

    hypochlorite. The ability to induce methaemoglobin formation was
    eliminated by treating the water by reverse osmosis followed by carbon
    filtration. Clearly, chloramine is capable of inducing
    methaemoglobinaemia at low concentrations when there is a large
    reservoir of chloramine. This is a decidedly different exposure
    pattern from that of normal humans and other mammals, as they consume
    small volumes of water relative to the volume of red blood cells that
    are exposed.

         Moore et al. (1980a) studied alterations of blood parameters in
    male A/J mice treated with 0, 2.5, 25, 50, 100 or 200 mg of
    monochloramine per litre in carbonate/bicarbonate-buffered (pH 8.9)
    drinking-water. Twelve animals were assigned to each group, and
    treatments were maintained over a 30-day period. Consistent with the
    interpretation provided above, there were no treatment-related effects
    on osmotic fragility, methaemoglobin levels, haemoglobin
    concentrations, reticulocyte counts or a number of other derived
    parameters. Haematocrits of mice treated with 50, 100 or 200 mg/litre
    were actually higher than those observed in control mice. White blood
    cell counts were not altered in these animals.

         Daniel et al. (1990a) conducted a more traditional 90-day study
    of monochloramine in Sprague-Dawley rats (10 animals per sex per
    dose). Treatment concentrations were 0, 25, 50, 100 or 200 mg/litre,
    corresponding to doses of 0, 1.8, 3.4, 5.8 or 9.0 mg/kg of body weight
    per day in males and 0, 2.6, 4.3, 7.7 or 12.1 mg/kg of body weight per
    day in females. Controls received carbonated, pH-adjusted
    drinking-water. A large number of haematological and clinical
    chemistry measures were included in the evaluation. Body weights were
    significantly depressed in both sexes at treatment concentrations in
    the 50-200 mg/litre range, but this appeared to be related to
    depressed water and food consumption. There were minor changes in
    organ to body weight ratios at the highest dose, but no evidence of
    treatment-related pathology was observed. Male rats were found to have
    decreased haematocrits at 100 mg/litre, and red blood cell counts were
    slightly depressed at 100 and 200 mg/litre. The authors concluded that
    monochloramine was more toxic than chlorine or chlorine dioxide.
    However, it must be noted that the changes in blood parameters were
    small and in themselves of no clinical significance. Other measures
    were not related to specific toxic reactions. Based on the decrease in
    organ and body weights observed in both sexes, the authors concluded
    that the no-observed-adverse-effect level (NOAEL) was 100 mg/litre,
    equivalent to 5.8 mg/kg of body weight per day.

         The work in rats was followed up with a second study in B6C3F1
    mice (Daniel et al., 1991a). Male and female B6C3F1 mice (10 per sex
    per group) were administered 0, 12.5, 25, 50, 100 or 200 mg of
    chloramine per litre of drinking-water for 90 days (calculated mean
    daily dose was 0, 2.5, 5.0, 8.6, 11.1 or 15.6 mg/kg of body weight for
    males and 0, 2.8, 5.3, 9.2, 12.9 or 15.8 mg/kg of body weight for
    females). Water consumption significantly decreased at 100 and 200
    mg/litre in males and at 25-200 mg/litre in females. Weight gain was
    significantly depressed in both sexes at 100 and 200 mg/litre.
    Neutrophil concentrations in blood were significantly depressed in

    both male and female mice at the two highest doses, but other white
    blood cell counts were unaltered. Absolute and relative spleen and
    liver weights were depressed at both 100 and 200 mg/litre. No gross or
    microscopic evidence of target organ toxicity was observed that could
    be related to treatment. Based on decreased organ weights, weight
    gain, and food and water consumption, the authors concluded that the
    NOAEL was 50 mg/litre, equivalent to 8.6 mg/kg of body weight per day.

         A 13-week study in which groups of 10 Sprague-Dawley rats were
    given drinking-water containing 200 mg of monochloramine per litre or
    buffered water as a control,  ad libitum or restricted to the same
    consumption as the monochloramine group, was designed to resolve some
    of the outstanding toxicological questions. The results of this study
    indicated that the reduced body weight gain and the minor biochemical,
    haematological, immunological and histopathological changes associated
    with exposure to 200 mg of monochloramine per litre (equivalent to
    21.6 mg/kg of body weight per day) in drinking-water were largely
    related to reduced water and food consumption (Poon et al., 1997).

         In a 9-week study, Exon et al. (1987) examined the ability of
    monochloramine to modify immunological parameters in male
    Sprague-Dawley rats (12 per dose) exposed to concentrations of 0, 9,
    19 or 38 mg of monochloramine per litre, equivalent to 0, 1.3, 2.6 or
    5.3 mg/kg of body weight per day. At the middle and highest dose,
    chloramine treatment was observed to increase prostaglandin E2
    synthesis by adherent resident peritoneal cells (which include
    macrophages) in response to lipopolysaccharide stimulation. No attempt
    was made to relate this finding to other indices of modified
    macrophage function. A small depression in spleen weights was observed
    at the highest dose. The implications of these data for immune
    function are not clear. Other measures of immune function did not
    reveal statistically significant changes with treatment. These
    included a decrease in antibody formation at the lowest and middle
    doses in response to keyhole limpet haemocyanin injection or
    delayed-type hypersensitivity reactions to bovine serum albumin
    injected into the footpad.

         The effect of monochloramine on skin irritation was tested by
    immersing Sencar mice into water containing chloramine at
    concentrations ranging from 1 to 1000 mg/litre for 10 min a day
    (Robinson et al., 1986). Unlike hypochlorous acid (pH 6.5) or
    hypochlorite (pH 8.5), chloramine did not produce hyperplasia of the
    skin.

    3.2.2  Reproductive and developmental toxicity

         Studies in laboratory animals have indicated no reproductive or
    developmental effects associated with chloramine. Abdel-Rahman et al.
    (1982a) administered monochloramine to female Sprague-Dawley rats at
    concentrations of 0, 1, 10 or 100 mg/litre (0, 0.15, 1.5 or 15 mg/kg
    of body weight per day) for 2.5 months prior to breeding and through
    gestation. Only six animals were assigned to each treatment group.
    Reproductive performance was comparable between groups, and fetal
    weights were not adversely affected by treatment. Between 50 and 60

    fetuses were available (male and female combined) for evaluation.
    There was no evidence of treatment-related skeletal or soft tissue
    anomalies.

         Carlton et al. (1986) examined the effects of monochloramine
    administered by gavage at doses of 0, 2.5, 5 or 10 mg/kg of body
    weight per day on the reproductive performance of Long-Evans rats.
    Males (12 per group) were treated from 56 days prior to and through
    mating, and females (24 per group) from 14 days prior to mating and
    throughout the mating period. No statistically significant effects on
    sperm morphology, concentration or motility were observed, nor were
    there any effects on fertility, viability, litter size, pup weights,
    day of eye opening or day of vaginal patency.

    3.2.3  Toxicity in humans

         The primary harmful effects of chloramine have been documented in
    humans poisoned by chloramine formed when household bleach was mixed
    with ammonia for use as a cleaning solution. Chloramine is a strong
    respiratory irritant. These effects were discussed in section 3.1.3.

         Forty-eight men completed an 8-week protocol during which diet
    and other factors known to affect lipid metabolism were controlled.
    During the first 4 weeks of the protocol, all subjects consumed
    distilled water. During the second 4 weeks, one-third of the subjects
    were assigned randomly to drink 1.5 litres of water containing 0, 2 or
    15 mg of monochloramine per litre each day. At 2 mg/litre, no
    significant changes were observed in total, HDL or LDL cholesterol,
    triglycerides or apolipoproteins A1, A2 or B. Parameters of thyroid
    function were unchanged. However, an increase in the level of
    apolipoprotein B was observed at 15 mg/litre (Wones et al., 1993b).

    3.2.4  Carcinogenicity and mutagenicity

         Shih & Lederberg (1976) first demonstrated that monochloramine
    induced mutation in a  Bacillus subtilis reversion assay. The
    concentration range studied extended from 18 to 74 mol/litre. A
    positive dose-response was observed through 56 mol/litre, but 74
    mol/litre was cytotoxic. Repair-deficient mutants of  B. subtilis, 
    rec3, recA and polyA5, were consistently more sensitive to the
    cytotoxic effects of chloramine, while the uvr and recB mutants were
    not. The sensitivity of the polyA5 mutants parallels the observations
    of Rosenkranz (1973) with sodium hypochlorite and suggests that DNA
    polymerase A is involved in the repair of DNA lesions produced by both
    chemicals. Thus, it is possible that a common intermediate or
    mechanism is involved in the mutagenic effects of hypochlorite and
    chloramine.

         A broader list of chloramines was tested by Thomas et al. (1987)
    in  Salmonella typhimurium tester strains TA97a, TA100 and TA102. The
    chloramines tested included those that could be formed at low levels
    from natural substrates in drinking-water or in the stomach. TA100 was
    found to consistently be the most sensitive strain. The most potent
    mutagens were the lipophilic dichloramines formed with histamine,

    ethanolamine and putrescine. The corresponding monochloramines were
    less potent. The more hydrophilic chloramines, such as
    taurine-chloramine, had little activity. Monochloramine was active in
    the 50 mol/litre range, remarkably consistent with the data of Shih &
    Lederberg (1976). Hypochlorous acid was inactive at all concentrations
    that were tested, up to and including concentrations that induced
    cytotoxicity.

         Ashby and co-workers (1987) were unable to induce clastogenic
    effects in the mouse bone marrow micronucleus assay when chloramine
    was administered orally. They suggested that the  in vitro 
    clastogenic effects were probably attributable to non-specific
    cytotoxic effects that are secondary to the release of hypochlorite to
    the media.

         Meier et al. (1985b) found that intraperitoneal administration of
    monochloramine to CD-1 mice at doses of up to 8 mg/kg of body weight
    was without significant effect on either micronuclei or chromosomal
    aberrations in the bone marrow. These data would appear to be
    consistent with the findings of Ashby et al. (1987).

         Studies on the carcinogenicity of chloramine are limited to a
    single set of 2-year experiments conducted by the NTP (1992).
    Drinking-water containing 0, 50, 100 or 200 mg of chloramine per litre
    was provided to F344 rats and B6C3F1 mice. Seventy animals of each
    species and of each sex within a species were assigned to each
    experimental and control group. Doses in rats were 0, 2.1, 4.8 or
    8.7 mg/kg of body weight per day in males and 0, 2.8, 5.3 or 9.5 mg/kg
    of body weight per day in females; doses in mice were 0, 5.0, 8.9 or
    15.9 mg/kg of body weight per day in males and 0, 4.9, 9.0 or
    17.2 mg/kg of body weight per day in females. Of some interest was the
    finding that two renal cell adenomas were found in male B6C3F1 mice
    treated with the high dose of chloramine. In addition, one renal
    adenoma was found in one male mouse treated with 100 mg/litre and in
    one female mouse treated with 200 mg/litre. While this tumour site is
    rare in both species, there was no real dose-response trend, nor were
    the differences between the control and treatment groups statistically
    significant. A second finding of some concern was an increase in the
    incidence of mononuclear cell leukaemia in F344 rats. This pathology
    was increased in rats treated with chloramine or hypochlorite,
    although the effects were not clearly dose-dependent. The incidence of
    mononuclear cell leukaemia was significantly greater than in
    concurrent controls and was elevated above the historical incidence as
    well. Nevertheless, these increases were not considered to be
    treatment-related. In part, this conclusion arose from the lack of a
    clear dose-response. It was also based on the fact that there was no
    comparable trend in male rats.

    3.2.5  Comparative pharmacokinetics and metabolism

         The pharmacokinetics of 36Cl derived from monochloramine have
    been examined in male Sprague-Dawley rats (Abdel-Rahman et al., 1983).
    These data are difficult to interpret because the specific form of the
    label is not known.

    3.3  Chlorine dioxide

    3.3.1  General toxicological properties and information on 
           dose-response in animals

         Despite its use as a disinfectant, there have been very few
    general toxicological evaluations of chlorine dioxide, because most
    studies have focused on its major by-product, chlorite, which is
    considered in section 4.6. The present review will first focus on the
    limited characterization of chlorine dioxide's general toxicology,
    then follow up with a discussion of its haematological and thyroid
    effects.

         Some very cursory investigations of chlorine dioxide's effects as
    a respiratory irritant were published by Haller & Northgraves (1955)
    in an article dealing with the general chemical properties of the
    compound. In essence, these data suggested that exposure to chlorine
    dioxide in air at a concentration of more than 420 mg/m3 (150 ppm)
    for longer than 15 min was fatal to guinea-pigs. The total study
    involved six guinea-pigs.

         Rats (3-5 per group) were exposed to chlorine dioxide in various
    concentrations (0.28-9520 mg/m3 [0.1-3400 ppm]) and for various
    periods (3 min-10 weeks). All the rats exposed daily to chlorine
    dioxide at 28 mg/m3 (10 ppm) died in less than 14 days. Purulent
    bronchitis and disseminated bronchopneumonia were found at necroscopy.
    No such changes were demonstrable in rats exposed to approximately
    0.28 mg/m3 (0.1 ppm) for about 10 weeks (Dalhamn, 1957). 

         The LC50 of chlorine dioxide in rats (5 per sex per group)
    exposed by inhalation for 4 h was 90 mg/m3 (32 ppm) (Ineris, 1996).

         Robinson et al. (1986) studied the ability of chlorine and
    alternative disinfectants to induce epidermal hyperplasia in the skin
    of Sencar mice. The thickness of the interfollicular epidermis was
    significantly increased by 10-min daily exposures to water containing
    up to 1000 mg of chlorine dioxide per litre for 4 days. The thickness
    of the epidermis was similar to that induced by an equivalent dose of
    hypochlorous acid. Unlike hypochlorous acid, however, there was no
    significant increase in skin thickness at concentrations of 300
    mg/litre or less.

         The study of Daniel et al. (1990a) was the first subchronic study
    that adhered to modern expectations of toxicological studies. These
    authors provided male and female Sprague-Dawley rats (10 per sex per
    treatment group) with 0, 25, 50, 100 or 200 mg of chlorine dioxide per
    litre of drinking-water for 90 days, equivalent to 0, 2, 4, 6 or 12
    and 0, 2, 5, 8 or 15 mg/kg of body weight per day for males and
    females, respectively. Conventional measures of body weight, organ
    weights, a broad battery of clinical chemistry parameters and
    histopathological examinations were all included in the study design.
    Body and organ weights were significantly depressed at 200 mg/litre in
    both sexes. This appeared to be secondary to depressed water
    consumption, which is known to be tightly coupled to food consumption

    in rats. The only significant histopathological damage found was
    goblet cell hyperplasia and inflammation. This was observed at all
    doses of chlorine dioxide in both male and female rats. Presumably
    this inflammation occurs as a result of volatilization of chlorine
    dioxide from the water bottle. The amount of chlorine dioxide actually
    inhaled as a result of volatilization from the drinking-water
    containing the lowest dose of chlorine dioxide (25 mg/litre) must be
    extremely low. This suggests that there might be some concern for
    sensitive individuals showering with water containing chlorine
    dioxide.

         The ability of chlorine dioxide to induce methaemoglobinaemia and
    haemolytic anaemia has received extensive study. Abdel-Rahman et al.
    (1980) found decreased red blood cell glutathione (GSH) concentrations
    and decreased osmotic fragility in Sprague-Dawley rats and white
    leghorn chickens given drinking-water containing chlorine dioxide
    concentrations of 1, 10, 100 and 1000 mg/litre for up to 4 months, but
    the changes were not consistently dose-related. However, the authors
    found that the morphology of red blood cells was modified (codocytes
    and echinocytes) in all dose groups, the severity increasing with
    increased treatment concentration. Methaemoglobin was not detected
    throughout these studies. However, there was no formal statistical
    evaluation of these results, and only four rats were assigned to each
    experimental group. Administration of acute doses of as little as 1
    mg/kg of body weight by gavage decreased red blood cell GSH
    concentrations. This response was not increased as dose was increased
    to 4 mg/kg of body weight.

         Abdel-Rahman and co-workers extended these observations to longer
    treatment periods in a subsequent publication (Abdel-Rahman et al.,
    1984a). Again, only four animals were assigned to each treatment
    group. At 7 and 9 months of treatment, red blood cells appeared to
    become resistant to osmotic shock at all treatment concentrations
    (1-1000 mg/litre). These data did not display a clear dose-response
    despite the large variation in the dose administered.

         The study of Abdel-Rahman et al. (1984a) also reported changes in
    the incorporation of 3H-thymidine into the DNA of various organs.
    Incorporation was significantly inhibited in testes and apparently
    increased in the intestinal mucosa. The effect on apparent DNA
    synthesis was particularly marked in the testes at 100 mg/litre,
    amounting to about 60% inhibition. These data are difficult to
    interpret for several reasons. First, rats were sacrificed 8 h after
    being injected with 3H-thymidine. Ordinarily, sacrifices are made
    30-60 min after injection because the blood is essentially depleted of
    3H-thymidine in an hour. Thus, it is not possible to determine if the
    lowered amount of label is related to decreased synthesis or to
    increased turnover of DNA. Second, the result was based on total
    counts in DNA, which makes it impossible to determine what cell type
    is affected or whether the change was associated with replicative or
    repair synthesis. Third, only four animals were used per experimental
    group.

         In a rising-dose protocol study, Bercz et al. (1982) evaluated
    the effects of chlorine dioxide on African green monkeys. These
    animals were provided chlorine dioxide in drinking-water at
    concentrations of 0, 30, 100 or 200 mg/litre, corresponding to doses
    of 0, 3.5, 9.5 or 11 mg/kg of body weight per day. Each dose was
    maintained for 30-60 days. Animals showed signs of dehydration at the
    highest dose (11 mg/kg of body weight per day), so exposure at that
    dose was discontinued. No effect was observed on any haematological
    parameter, including methaemoglobinaemia. However, statistically
    significant depressions in serum thyroxine levels were observed when
    animals were dosed with chlorine dioxide at a concentration of
    100 mg/litre. No effect had been observed in a prior exposure of the
    same animals to 30 mg/litre for 30 days. The NOAEL in this study was
    3.5 mg/kg of body weight per day.

         The effects of chlorine dioxide on thyroid function were followed
    up by Harrington et al. (1986). Thyroxine levels in African green
    monkeys administered drinking-water containing 100 mg of chlorine
    dioxide per litre (4.6 mg/kg of body weight per day) were again found
    to be depressed at 4 weeks of treatment, but rebounded to above-normal
    levels after 8 weeks of treatment. These investigators also found
    significantly depressed thyroxine levels in rats treated with 100 or
    200 mg of chlorine dioxide per litre in drinking-water (equivalent to
    14 and 28 mg/kg of body weight per day) for 8 weeks. This change was
    dose-related. Lower doses were not examined in the rat study. The
    authors indicated that the results were based on 12 determinations; it
    was not clear if these measurements were made on individual animals. 

         A set of  in vivo and  in vitro experiments was conducted in an
    attempt to explain the effects of chlorine dioxide on serum thyroid
    hormone concentrations. The authors demonstrated that chlorine dioxide
    oxidizes iodide to reactive iodine species that would bind to the
    stomach and oesophageal epithelium. Rat chow that was treated with
    chlorine dioxide at approximately 80 mg/litre was found to increase
    the binding of iodine to chow constituents. This activation of iodine
    resulted in retention of labelled iodine in the ileum and colon and
    reduced uptake by the thyroid gland. Previous work demonstrated that
    chlorine dioxide was more effective than chlorine in activating iodide
    to a form that would covalently bind with a variety of natural
    foodstuffs (Bercz et al., 1986). Based on these observations, the
    authors concluded that the effects of chlorine dioxide were probably
    due to altered gastrointestinal absorption of iodide and reduced
    uptake into the thyroid gland.

    3.3.2  Reproductive and developmental toxicity

         A number of reproductive effects have been reported in studies
    with laboratory animals, but the relevance for humans of these
    findings remains uncertain. The reproductive effects of chlorine
    dioxide in Long-Evans rats were studied by Carlton et al. (1991).
    Chlorine dioxide was administered by gavage at doses of 2.5, 5 or 10
    mg/kg of body weight per day to male rats (12 per group) for 56 days
    prior to and through mating and to female rats (24 per group) from 14
    days prior to mating and through pregnancy. Fertility measures were

    not significantly different among the dose groups. There were no
    dose-related changes in sperm parameters (i.e., concentration,
    motility, progressive movement or morphology). Thyroid hormone levels
    were altered significantly, but not in a consistent pattern. The only
    significant difference was significantly depressed vaginal weights in
    female pups whose dams had been treated with 10 mg/kg of body weight
    per day.

         An evaluation of the effects of chlorine dioxide on the fetal
    development of Sprague-Dawley rats was conducted by Suh et al. (1983).
    Chlorine dioxide was administered at 0, 1, 10 or 100 mg/litre (0, 0.1,
    1 or 10 mg/kg of body weight per day) for 2.5 months prior to mating
    and throughout gestation. The total number of implants per dam was
    significantly reduced at the highest concentration of chlorine
    dioxide. The percentage of anomalous fetuses was increased in a
    dose-related manner, but the response was not statistically
    significant. These anomalies arose primarily as the percentage of
    abnormal or incomplete sternebrae in treated rats relative to
    controls. The lack of statistical significance was undoubtedly related
    to the relatively few female rats that were included in the study (6-8
    females per treatment group). As a consequence, the results of this
    study must be considered inconclusive.

         Orme et al. (1985) found that chlorine dioxide administered in
    the drinking-water of female Sprague-Dawley rats (13-16 per dose) at
    concentrations of 0, 2, 20 or 100 mg/litre (0, 1, 3 or 14 mg/kg of
    body weight per day) throughout pregnancy and through weaning
    decreased thyroxine levels in the serum of the pups at 100 mg/litre.
    This was associated with delayed development of exploratory behaviour
    in the pups away from their dams, and this, in turn, was probably due
    to an indirect effect on iodine uptake. In a second experiment, pups
    given 14 mg of chlorine dioxide per kg of body weight per day directly
    by gavage on postnatal days 5-20 showed significantly depressed
    activity and a decrease in serum thyroxine levels. Studies by the same
    group (Taylor & Pfohl, 1985) indicated that cerebellar and forebrain
    cell counts (based on DNA measurements) were depressed in 11-day-old
    pups that had been treated with chlorine dioxide at 14 mg/kg of body
    weight per day by gavage from 5 days of age. Cerebellar cell counts
    remained depressed in rats at 21 days, but forebrain counts were
    essentially the same as in controls. At 50-60 days of age, the
    locomotor activity (measured by wheel-running) of these animals was
    depressed relative to control animals.

         The effects of chlorine dioxide on brain development were
    examined further by Toth et al. (1990). These authors administered
    chlorine dioxide by gavage at 14 mg/kg of body weight per day from
    postnatal day 1 to 20. Body weight was reduced, but cerebellar weight
    was unaltered at any age. Forebrain weight and protein content were
    reduced on postnatal days 21 and 35. DNA content was depressed on
    postnatal day 35, and the number of dendritic spines on cerebral
    cortical pyramidal cells was significantly reduced. No
    histopathological changes in the forebrain, cerebellum or brain stem
    were observed. There were no consistent changes in serum thyroxine or
    triiodothyronine levels in treated animals.

         Collectively, these data suggest some effects of chlorine dioxide
    on brain development. In most studies, there are suggestions of
    modified thyroid function associated with these effects. It must be
    pointed out that the changes in thyroid hormone levels are modest,
    much less than are produced with classical antithyroid drugs such as
    propylthiouracil (Toth et al., 1990).

    3.3.3  Toxicity in humans

         The effects of chlorine dioxide were assessed in a two-phase
    study in 10 healthy male volunteers. The first study was a rising-dose
    tolerance study (Lubbers & Bianchine, 1984) in which doses of chlorine
    dioxide were increased from 0.1 to 24 mg/litre, administered in two
    500-ml portions. The maximum dose for a 70-kg person was 0.34 mg/kg of
    body weight. The details of this study were described in section
    3.1.3. Some small changes in a variety of clinical chemistry
    parameters were observed, but none was found to be outside the
    accepted range of normal. The second phase of the experiment involved
    the daily administration of a 500-ml portion of a solution containing
    5 mg/litre to 10 healthy volunteers for a period of 12 weeks (Lubbers
    et al., 1984a). Again, measurement of a large battery of clinical
    chemistry parameters and routine physical examination failed to
    identify any effects of chlorine dioxide that fell outside of the
    normal range. Parameters yielding significant differences appeared to
    be primarily a result of parallel drift of values with the control
    group.

         As with other disinfectants, it is important to recognize that
    chlorine dioxide is a potent respiratory irritant. No quantitative
    data can be used to construct a dose-response relationship for this
    effect. 

    3.3.4  Carcinogenicity and mutagenicity

         The mutagenic or clastogenic effects of chlorine dioxide have
    received little attention. Ishidate et al. (1984) found chlorine
    dioxide to be positive in  Salmonella typhimurium tester strain
    TA100. A linear dose-response was observed at concentrations between 2
    and 20 g per plate. Chlorine dioxide was ineffective as a clastogenic
    agent in a CHO system.

         Meier et al. (1985b) evaluated the ability of chlorine dioxide to
    induce chromosomal aberrations and micronuclei in bone marrow of CD-1
    mice or sperm head anomalies. Chlorine dioxide failed to produce such
    damage following gavage doses of up to 16 mg/kg of body weight for 5
    days.

         With the exception of a 1949 study by Haag (cited in TERA, 1998),
    which has serious limitations, no tests of the carcinogenic activity
    of chlorine dioxide in experimental animals were identified in the
    scientific literature.

    3.3.5  Comparative pharmacokinetics and metabolism

         There are significant differences in the pharmacokinetics of
    36Cl obtained from different disinfectants. The absorption rate for
    36Cl-labelled chlorine dioxide was at least 10 times that observed
    with chlorine, chloramine or chloride. The relative amount of 36Cl
    that is eliminated in the urine and faeces at 24 h has a distinct
    pattern from that observed with other disinfectants and sodium
    chloride. However, the terminal half-life of the 36Cl appears similar
    for all disinfectants. These data suggest that the form of 36Cl that
    is being absorbed differs chemically with the different disinfectants.
    In the case of chlorine dioxide, this is supported by the observation
    that measurable amounts of chlorite (about 3% of the original dose)
    are eliminated in the urine during the first 24 h, and chlorite
    comprises about 20% of the label present in blood 72 h after
    administration of the test dose (Abdel-Rahman et al., 1980). However,
    this higher absorption rate is not explained by the absorption rates
    of chlorite and chlorate, which are about one-tenth as rapid
    (Abdel-Rahman et al., 1982b). This suggests that some of the
    absorption could be as chlorine dioxide itself. On the surface, this
    hypothesis would seem to be incompatible with the high reactivity of
    this disinfectant. As with other disinfectants, the terminal
    elimination phases observed for 36Cl from chlorine dioxide seem
    compatible with the hypothesis that the bulk of the elimination is as
    chloride ion.

    4.  TOXICOLOGY OF DISINFECTANT BY-PRODUCTS

    4.1  Trihalomethanes

         As a class, the THMs are generally the most prevalent by-products
    of drinking-water disinfection by chlorine. A variety of
    non-neoplastic toxic effects have been associated with short-term and
    long-term exposure of experimental animals to high doses of THMs, and
    each of the four most common THMs -- chloroform, BDCM, DBCM and
    bromoform -- has been shown to be carcinogenic to rodents in high-dose
    chronic studies. Chloroform is generally the predominant THM in
    chlorinated water and is also the most extensively studied chemical of
    this class. Because the World Health Organization (WHO) recently
    published an Environmental Health Criteria monograph on chloroform
    (IPCS, 1994), this section will only update the information contained
    in that publication. A thorough review of findings relevant to the
    toxicology of the brominated THMs is included.

         As with the other DBPs, the chemical and physical properties of
    the THMs influence their potential routes of human exposure, their
    pharmacokinetic behaviour, their toxicity and methods for conducting
    toxicological studies with these compounds. The THMs are volatile
    liquids at room temperature; therefore, as these chemicals vaporize
    during water usage (e.g., showering), inhalation becomes an important
    exposure route in addition to ingestion. Volatility decreases somewhat
    with bromine substitution, but each of the brominated THMs evaporates
    from drinking-water. Like that of other alkanes, the water solubility
    of THMs is poor, although adequate to permit dissolution of the low
    levels generated via water disinfection. When administered at higher
    levels to animals in toxicity experiments, THMs are often either
    emulsified in aqueous solutions or dissolved in oils. The use of oils
    as vehicles of administration can significantly alter the
    pharmacokinetics and toxicity of the THMs. Bromine substitution
    enhances the lipid solubility of the halomethanes (and, consequently,
    uptake into tissues) and generally increases their chemical reactivity
    and the likelihood of biotransformation to a reactive intermediate.
    Because toxicity is dependent upon a reactive compound actually
    reaching a sensitive target site, greater bromine substitution may not
    necessarily translate into greater  in vivo toxicity (i.e., innocuous
    reactions may occur, preventing arrival at target sites). Perhaps to a
    greater extent than with other chemicals in this class, BDCM appears
    to reach a variety of target tissues where it can be readily
    metabolized to several intermediates, leading to adverse effects in
    experimental animals.

    4.1.1  Chloroform

         In 1994, the WHO published an Environmental Health Criteria
    monograph on chloroform (IPCS, 1994). The following sections will
    update that document with the most recent findings from health-related
    chloroform research. Another evaluation of chloroform was included in
    the 1998 Addendum to the WHO  Guidelines for drinking-water quality 
    (WHO, 1998).

    4.1.1.1  General toxicological properties and information on 
             dose-response in animals

    1)   Acute toxicity

         Keegan et al. (1998) determined the
    lowest-observed-adverse-effect level (LOAEL) and NOAEL for the
    induction of acute hepatotoxicity following oral administration of
    chloroform in an aqueous vehicle to male F344 rats. Based on
    elevations of serum clinical chemistry indicators of liver damage, a
    LOAEL of 0.5 mmol/kg of body weight (60 mg/kg of body weight) and a
    NOAEL of 0.25 mmol/kg of body weight (30 mg/kg of body weight) were
    established. In a corn oil gavage study of single-dose chloroform
    effects, an increase in renal cell proliferation was observed at doses
    as low as 10 mg/kg of body weight in male Osborne-Mendel rats and
    90 mg/kg of body weight in male F344 rats (Templin et al., 1996a). The
    only increase in the hepatic labelling index was in F344 rats given
    477 mg/kg of body weight. Effects in the nasal passages of both rat
    strains at 90 mg/kg of body weight and above included oedema and
    periosteal hypercellularity. Gemma et al. (1996) dosed male B6C3F1
    mice with chloroform (150 mg/kg of body weight) by gavage and observed
    increases in cell proliferation in both the liver and kidneys. The
    effect was more dramatic in the kidneys, where severe necrosis was
    also noted.

         Nephrotoxicity of chloroform was evaluated in male Sprague-Dawley
    rats treated orally with single doses of chloroform using corn oil or
    an aqueous preparation (5%) of Emulphor or Tween 85 as vehicle (10
    ml/kg of body weight). Comparison between gavage vehicles indicated
    clear trends for enhanced potency and severity of nephrotoxic effects
    with corn oil administration of chloroform (Raymond & Plaa, 1997).

    2)   Short-term toxicity

         Chloroform was administered by corn oil gavage to male B6C3F1
    mice at doses of 0, 34, 90, 138 or 277 mg/kg of body weight for 4 days
    or 3 weeks (5 days per week) (Larson et al., 1994a). Mild degenerative
    changes in centrilobular hepatocytes were noted in mice given 34 and
    90 mg/kg of body weight per day after 4 days of treatment, but these
    effects were absent at 3 weeks. At 138 and 277 mg/kg of body weight
    per day, centrilobular necrosis was observed at 4 days and with
    increased severity at 3 weeks. Hepatic cell proliferation was
    increased in a dose-dependent manner at all chloroform doses after 4
    days, but only in the 277 mg/kg of body weight dose group at 3 weeks.
    Renal tubular necrosis was observed in all dose groups after 4 days,
    while 3 weeks of exposure produced severe nephropathy at the highest
    dose and regenerating tubules at the lower doses. The nuclear
    labelling index was increased in the proximal tubules at all doses
    after 4 days of treatment, but was elevated only in the two highest
    dose groups after 3 weeks.

         In a similar study (Larson et al., 1994b), female B6C3F1 mice
    were administered chloroform dissolved in corn oil by gavage at doses
    of 0, 3, 10, 34, 238 or 477 mg/kg of body weight per day for 4 days or
    3 weeks (5 days per week). Dose-dependent changes included
    centrilobular necrosis and markedly elevated labelling index in mice
    given 238 and 477 mg/kg of body weight per day. The NOAEL for
    histopathological changes was 10 mg/kg of body weight per day, and for
    induced cell proliferation, 34 mg/kg of body weight per day.

         In an inhalation study, Templin et al. (1996b) exposed BDF1 mice
    to chloroform vapour at concentrations of 0, 149 or 446 mg/m3 (0, 30
    or 90 ppm) 6 h per day for 4 days or 2 weeks (5 days per week). In the
    kidneys of male mice exposed to 149 and 446 mg/m3 (30 and 90 ppm),
    degenerative lesions and 7- to 10-fold increases in cell proliferation
    were observed. Liver damage and an increased hepatic labelling index
    were noted in male mice exposed to 149 and 446 mg/m3 (30 and 90 ppm)
    and in female mice exposed to 446 mg/m3 (90 ppm). Both doses were
    lethal in groups exposed for 2 weeks (40% and 80% mortality at 149 and
    446 mg/m3 [30 and 90 ppm], respectively).

         Female F344 rats were given chloroform by corn oil gavage for 4
    consecutive days or 5 days per week for 3 weeks (Larson et al.,
    1995b). In the liver, mild degenerative centrilobular changes and
    dose-dependent increases in hepatocyte proliferation were noted at
    doses of 100, 200 and 400 mg/kg of body weight per day. At 200 and
    400 mg/kg of body weight per day, degeneration and necrosis of the
    renal cortical proximal tubules were observed. Increased regenerative
    proliferation of epithelial cells lining proximal tubules was seen at
    doses of 100 mg/kg of body weight per day or more. Lesions of the
    olfactory mucosa lining the ethmoid region of the nose (new bone
    formation, periosteal hypercellularity and increased cell replication)
    were seen at all doses, including the lowest dose of 34 mg/kg of body
    weight per day. Larson et al. (1995a) also administered chloroform to
    male F344 rats by corn oil gavage (0, 10, 34, 90 or 180 mg/kg of body
    weight per day) or in the drinking-water (0, 60, 200, 400, 900 or
    1800 mg/litre) for 4 days or 3 weeks. Gavage of 90 or 180 mg/kg of
    body weight per day for 4 days induced mild to moderate degeneration
    of renal proximal tubules and centrilobular hepatocytes -- changes
    that were no longer present after 3 weeks. Increased cell
    proliferation in the kidney was noted only at the highest gavage dose
    after 4 days. The labelling index was elevated in the livers of the
    high-dose group at both time points. With drinking-water
    administration, rats consuming the water containing 1800 mg/litre were
    dosed at a rate of 106 mg/kg of body weight per day, but no increase
    in renal or hepatic cell proliferation was observed at this or any
    lower dose.

         In a study carried out to evaluate whether exposure to chloroform
    in drinking-water would interact with the activity of chloroform when
    administered by gavage in corn oil, female B6C3F1 mice were exposed
    to chloroform in drinking-water for 33 days at 0, 300 or 1800 mg/litre
    or for 31 days at 0, 120, 240 or 480 mg/litre. Three days prior to
    termination, mice also received a daily dose of 263 mg of chloroform
    per kg of body weight per day, administered by gavage in corn oil.

    Exposure to chloroform in drinking-water reduced both the
    hepatotoxicity and the enhanced cell proliferation elicited in
    response to chloroform administered by gavage in corn oil (Pereira &
    Grothaus, 1997).

         The cardiotoxicity of chloroform was examined in male Wistar rats
    given daily doses of 37 mg/kg of body weight (0.31 mmol/kg) by gavage
    in olive oil for 4 weeks (Muller et al., 1997). Chloroform caused
    arrhythmogenic and negative chronotropic and dromotropic effects as
    well as extension of the atrioventricular conduction time and
    depressed myocardial contractility.

         A 90-day chloroform inhalation study was conducted using male and
    female B6C3F1 mice and exposure concentrations of 0, 1.5, 10, 50, 149
    and 446 mg/m3 (0, 0.3, 2, 10, 30 and 90 ppm) for 6 h per day, 7 days
    per week (Larson et al., 1996). Large, sustained increases in
    hepatocyte proliferation were seen in the 446 mg/m3 (90 ppm) groups
    at all time points (4 days and 3, 6 and 13 weeks). In the more
    sensitive female mice, a NOAEL of 50 mg/m3 (10 ppm) for this effect
    was established. Renal histopathology and regenerative hyperplasia
    were noted in male mice at 50, 149 and 446 mg/m3 (10, 30 and 90 ppm).
    In another 90-day inhalation study, F344 rats were exposed to
    chloroform at concentrations of 0, 10, 50, 149, 446 or 1490 mg/m3 (0,
    2, 10, 30, 90 or 300 ppm) for 6 h per day, 7 days per week. The 1490
    mg/m3 (300 ppm) level was extremely toxic and deemed by the authors
    to be inappropriate for chronic studies. Increases in renal epithelial
    cell proliferation in cortical proximal tubules were observed at
    concentrations of 149 mg/m3 (30 ppm) and above. Hepatic lesions and
    increased proliferation were noted only at the highest exposure level.
    In the ethmoid turbinates of the nose, enhanced bone growth and
    hypercellularity in the lamina propria were observed at concentrations
    of 50 mg/m3 (10 ppm) and above, and a generalized atrophy of the
    turbinates was seen at all exposure levels after 90 days (Templin et
    al., 1996c).

    3)   Reproductive and developmental toxicity

         Rat embryo culture was used to assess the developmental effects
    of chloroform (Brown-Woodman et al., 1998). The effect and
    no-effect culture medium concentrations of chloroform were 2.06 and
    1.05 mol/ml. The authors estimated that fatal or near-fatal blood
    levels would be required in the mother for the embryotoxic level to be
    reached.

    4.1.1.2  Toxicity in humans

         Fatal acute chloroform intoxication via inhalation was reported
    to cause cardiomyocyte fragmentation and waviness indicative of acute
    heart failure possibly caused by arrhythmias or cardiac depression
    (Harada et al., 1997). These observations are consistent with the
    results of the short-term rat study (Templin et al., 1996b) described
    above in section 4.1.1.1.

    4.1.1.3  Carcinogenicity and mutagenicity

         Jamison et al. (1996) reported that F344 rats exposed to a high
    concentration of chloroform vapour (1490 mg/m3 [300 ppm]) for 90 days
    developed atypical glandular structures lined by intestinal-like
    epithelium and surrounded by dense connective tissue in their livers.
    These lesions appeared to arise from a population of cells remote from
    the bile ducts. The authors also observed a treatment-related increase
    in transforming growth factor-alpha (TGF-alpha) immunoreactivity in
    hepatocytes, bile duct epithelium, bile canaliculi and oval cells and
    an increase in transforming growth factor-beta (TGF-)
    immunoreactivity in hepatocytes, bile duct epithelium and intestinal
    crypt-like ducts. The lesions occurred only in conjunction with
    significant hepatocyte necrosis, regenerative cell proliferation and
    increased growth factor expression or uptake.

         Chloroform was tested for mutagenicity and clastogenicity by Le
    Curieux et al. (1995) and was negative in the SOS chromotest, the Ames
    fluctuation test and the newt micronucleus test. It appeared to these
    authors that the presence of bromine substituents was needed for
    genotoxic activity in the THM class. Pegram et al. (1997) examined
    chloroform mutagenicity in a strain of  Salmonella typhimurium TA1535
    transfected with rat glutathione- S-transferase (GST) T1-1.
    Chloroform was negative in this assay over a range of concentrations
    (992-23 800 mg/m3 [200-4800 ppm]) that produced large dose-dependent
    increases in revertants with BDCM. A doubling of revertants was
    induced by chloroform in the GST-transfected strain only at the two
    highest concentrations tested (95 200 and 127 000 mg/m3 [19 200 and
    25 600 ppm]). Brennan & Schiestl (1998) found that chloroform induced
    intrachromosomal recombination in the yeast strain  Saccharomyces 
     cerevisiae at culture medium concentrations of 3-5.6 mg/ml.

         In an  in vivo study, Potter et al. (1996) found that chloroform
    did not induce DNA strand breaks in the kidneys of male F344 rats
    following seven daily doses of 1.5 mmol/kg of body weight. In
    long-term mutagenicity studies with chloroform in female  lacI 
    transgenic B6C3F1 mice conducted by Butterworth et al. (1998), the
    mice were exposed daily by inhalation to chloroform concentrations of
    0, 50, 149 or 446 mg/m3 (0, 10, 30 or 90 ppm) for 6 h per day, 7 days
    per week, and  lacI mutant frequency was determined after 10, 30, 90
    and 180 days of exposure. No increase in  lacI mutant frequency was
    observed in the liver at any dose or time point with chloroform.

    4.1.1.4  Comparative pharmacokinetics and metabolism

         The percutaneous absorption of 14C-chloroform was examined
     in vivo using human volunteers and  in vitro using fresh, excised
    human skin in a flow-through diffusion cell system (Dick et al.,
    1995). Aqueous and ethanol solutions of chloroform were applied to the
    forearm (16 and 81 g/cm) of volunteers, and absorption was determined
    to be 7.8% from the water vehicle and 1.6% from ethanol. More than 95%
    of the absorbed dose was excreted via the lungs (88% as carbon
    dioxide), and maximum pulmonary excretion occurred between 15 min and

    2 h after dosing.  In vitro, 5.6% of a low dose and 7.1% of a high
    dose were absorbed (skin plus perfusate).

         The systemic uptake of chloroform during dermal exposure was also
    studied in hairless rats (Islam et al., 1996). Animals were immersed
    in water containing chloroform for 30 min, and the compound was
    detected in blood as early as 4 min following exposure. About 10 mg of
    chloroform were systemically absorbed after dermal exposure of a rat
    to an aqueous solution of 0.44 mg/ml.

         Absorption and tissue dosimetry of chloroform were evaluated
    after gavage administration in various vehicles to male Fischer 344
    rats and female B6C3F1 mice (Dix et al., 1997). Animals received a
    single dose of chloroform (15-180 and 70-477 mg/kg of body weight for
    rats and mice, respectively) in corn oil, water or aqueous 2% Emulphor
    (dose volumes of 2 and 10 ml/kg of body weight for rats and mice,
    respectively). Blood, liver and kidney chloroform concentration-time
    courses were determined. Gavage vehicle had minimal effects on
    chloroform dosimetry in rats. In mice, however, tissue chloroform
    concentrations were consistently greater for aqueous versus corn oil
    vehicle. At the low vehicle volume used for rats (2 ml/kg of body
    weight), gavage vehicle may not play a significant role in chloroform
    absorption and tissue dosimetry; at the higher vehicle volume used for
    mice (10 ml/kg of body weight), however, vehicle may be a critical
    factor.

         Because chloroform metabolism was reviewed in detail in the
    recent Environmental Health Criteria monograph (IPCS, 1994), the
    primary discussion of the comparative metabolism of the THMs as a
    class can be found in section 4.1.2.6 of the present report.

         The contributions of cytochromes P450 (CYP) 2E1 and 2B1/2 to
    chloroform hepatotoxicity were investigated in male Wistar rats
    (Nakajima et al., 1995). The severity of toxicity observed in
    differentially induced rats suggests that CYP2E1 is a low
    Michaelis-Menten constant ( Km) isoform and CYP2B1/2 is a high
     Km isoform for chloroform activation. A high dose of chloroform
    (0.5 ml/kg of body weight) induced CYP2E1 but decreased CYP2B1/2.
    Testai et al. (1996) generated similar results in a study examining
    the involvement of these isozymes in  in vitro chloroform metabolism.
    At a low substrate concentration (0.1 mmol/litre), oxidative
    metabolism by liver microsomes was dependent primarily on CYP2E1; at 5
    mmol/litre, on the other hand, CYP2B1/2 was the major participant
    responsible for chloroform activation, although CYP2E1 and CYP2C11
    were also significantly involved. The reductive pathway was expressed
    only at 5 mmol/litre and was not significantly increased by any CYP
    inducer tested.

         The reductive metabolism of chloroform by rat liver microsomes
    was examined by Testai et al. (1995). In hypoxic (1% oxygen partial
    pressure) and anoxic (0% oxygen partial pressure) incubations using
    microsomes from phenobarbital-induced animals, no evidence of
    formation of monochloromethyl carbene was found. Dichloromethane was
    detected as a metabolite of chloroform under variable oxygenation

    conditions using microsomes from phenobarbital-induced animals. With
    uninduced microsomes, significant levels of dichloromethane were
    formed only in hypoxic or anoxic incubations. In an  in vivo study of
    chloroform reductive metabolism, Knecht & Mason (1991) detected no
    free radical adducts in the bile of rats treated with chloroform,
    while radicals were detected from bromoform. Lipid adducts resulting
    from the reductive metabolism of chloroform by hepatocytes appeared to
    be generated by the unspecific attack of the radical on the
    phospholipid fatty acyl chains (Guastadisegni et al., 1996). The
    primary lipid adduct has now been identified as a modified
    phosphatidylethanolamine, with the phosgene-derived carbonyl bound to
    the amine of the head group (Guastadisegni et al., 1998). Waller &
    McKinney (1993) found that chloroform had a lower theoretical
    potential to undergo reductive metabolism than the brominated THMs.
    Ade et al. (1994) reported that microsomes from the renal cortex of
    DBA/2J mice can metabolize chloroform through the reductive and
    oxidative pathways, as had been previously described using hepatic
    microsomes. However, cytolethality of chloroform to freshly isolated
    rodent hepatocytes was not increased under reduced oxygen tension,
    indicating that reductive metabolism does not contribute to
    chloroform-induced toxicity (Ammann et al., 1998).

         The potential of chloroform to participate in the recently
    discovered GSH conjugation pathway for brominated THMs has been
    investigated (Pegram et al., 1997). The GST examined in this study has
    a very low affinity for chloroform compared with the brominated THMs.
    Chloroform conjugation with GSH occurred only at extremely high
    substrate concentrations.

    4.1.1.5  Mode of action

         Direct DNA reactivity and mutagenicity cannot be considered to be
    key factors in chloroform-induced carcinogenesis in experimental
    animals. A substantial body of data demonstrates a lack of direct
     in vivo or  in vitro genotoxicity of chloroform. If THMs produce
    their genotoxic effects primarily via the GSH conjugation mechanism,
    the results of Pegram et al. (1997) indicate that chloroform would be
    mutagenic in mammals only at lethal doses.

         There is, however, compelling evidence to support a mode of
    action for tumour induction based on metabolism of chloroform by the
    target cell population, followed by cytotoxicity of oxidative
    metabolites and regenerative cell proliferation. The evidence for the
    link with cytotoxicity is strongest for hepatic and renal tumours in
    the mouse and more limited for renal tumours in the rat (ILSI, 1997).
    A number of recent studies support the hypothesis that chloroform acts
    to produce cancer in rodents through a non-genotoxic/cytotoxic mode of
    action, with carcinogenesis resulting from events secondary to
    chloroform-induced cytolethality and regenerative cell proliferation
    (Larson et al., 1994a,b, 1996; Pereira, 1994; Templin et al.,
    1996a,b,c, 1998). These studies have shown that organ toxicity and
    regenerative hyperplasia are associated with the tumorigenicity of
    chloroform and are apparently the key steps in its carcinogenic mode
    of action. Thus, sustained toxicity would result in tumour

    development. Chloroform induces liver and kidney tumours in long-term
    rodent cancer bioassays only at doses that induce frank cytotoxicity
    in these target organs. Furthermore, there are no instances of
    chloroform-induced tumours that are not preceded by this pattern of
    dose-dependent toxic responses (Golden et al., 1997).

         The organ toxicity and carcinogenicity of chloroform are
    dependent on oxidative metabolism and levels of CYP2E1. Numerous
    studies have also shown that oxidative metabolism by CYP2E1 generates
    highly reactive metabolites (phosgene and hydrogen chloride), which
    would lead to cytotoxicity and regenerative hyperplasia. 

    4.1.2  Bromodichloromethane

    4.1.2.1  General toxicological properties and information on 
             dose-response in animals

    1)   Acute toxicity

         The acute oral lethality of the brominated THMs has been
    determined in ICR Swiss mice (Bowman et al., 1978) and Sprague-Dawley
    rats (Chu et al., 1980). The resulting LD50s for BDCM were 450 and
    900 mg/kg of body weight for male and female mice, respectively, and
    916 and 969 mg/kg of body weight for male and female rats,
    respectively. Clinical observations of animals dosed with high levels
    of BDCM in these studies and others (NTP, 1987) included ataxia,
    sedation, laboured breathing and anaesthesia (500 mg/kg of body
    weight), as well as gross evidence of liver and kidney damage. Hewitt
    et al. (1983) gave single doses of BDCM to male Sprague-Dawley rats by
    corn oil gavage and found doses of 1980 mg/kg of body weight and above
    to be lethal. Little clinical evidence of hepatic or renal toxicity
    was observed at doses below 1980 mg/kg of body weight.

         Acute hepatotoxic and nephrotoxic responses to orally dosed BDCM
    and various factors affecting these toxicities (e.g., gavage vehicle
    and glutathione status) have been studied in male F344 rats. Lilly et
    al. (1994, 1997a) examined the time course of toxicity and
    dose-response relationships following oral administration of aqueous
    solutions of BDCM. BDCM-induced liver toxicity was maximal at 24 h
    after dosing with 1-3 mmol/kg of body weight (164-492 mg/kg of body
    weight), as indicated by elevations in serum levels of aspartate
    aminotransferase (ASAT), alanine aminotransferase (ALAT), sorbitol
    dehydrogenase (SDH) and lactate dehydrogenase (LDH) and
    histopathological observations of centrilobular vacuolar degeneration
    and hepatocellular necrosis. Significant abatement of hepatic toxicity
    was noted by 48 h post-dosing. The acute oral NOAEL and LOAEL for
    liver toxicity following aqueous delivery of BDCM, based on elevations
    in serum enzymes, were determined to be 0.25 and 0.5 mmol of BDCM per
    kg of body weight (41 and 82 mg/kg of body weight), respectively
    (Keegan et al., 1998). BDCM and chloroform appear to be equipotent
    hepatotoxicants in rats at 24 h after exposure, but BDCM causes more
    persistent damage to the liver, based on observations at 48 h
    post-dosing (Lilly et al., 1997a; Keegan et al., 1998).

         Kidney toxicity after corn oil or aqueous dosing of BDCM
    (1.5-3 mol/kg of body weight) peaked between 24 and 48 h, as indicated
    by elevations in kidney weight, urinary  N-acetyl--glucosaminidase,
    ASAT, ALAT, LDH and protein, serum urea and creatinine, and
    histopathological findings of renal tubule degeneration and necrosis
    (Lilly et al., 1994, 1997a). The actual time of peak renal effects was
    dose-dependent; in contrast to findings in the liver, toxicity was
    increasingly prolonged in the kidney with increasing dose.
    Nephrotoxicity has been noted in rats given single BDCM doses as low
    as 200 mg or 1.2 mmol/kg of body weight (Lilly et al., 1994, 1997a),
    and BDCM is a slightly more potent acute oral renal toxicant than
    chloroform (based on the magnitude of the responses), especially at
    lower doses (Lilly et al., 1997a). Kroll et al. (1994a,b) found that
    among the THMs that occur in drinking-water, BDCM was the most potent
    inducer of renal dysfunction in rats following intraperitoneal
    injection of single 3 mmol/kg of body weight doses. Glomerular
    filtration, renal concentrating ability, and proximal tubular
    secretion and reabsorption were all more severely affected by BDCM
    than by chloroform.

         Several factors have been found to influence dose-response
    relationships for BDCM toxicity. Acute hepatotoxicity and
    nephrotoxicity were more severe after administration of 400 mg of BDCM
    per kg of body weight in corn oil than when the same dose was given in
    an aqueous vehicle (Lilly et al., 1994). However, vehicle differences
    at a lower dose (200 mg/kg of body weight), although less pronounced,
    were reversed: greater renal toxicity at this dose was associated with
    the aqueous vehicle. The adverse renal and hepatic effects of BDCM
    were also exacerbated in GSH-depleted rats (Gao et al., 1996) and in
    rats that were dosed during the active period of their diurnal cycle
    (Pegram et al., 1993). Induction of the cytochrome P-450 isozymes
    CYP2E1 and CYP2B1/2 also potentiated acute liver toxicity, but not
    renal toxicity, following dosing with BDCM (Thornton-Mannning et al.,
    1994).

    2)   Short-term toxicity

         Studies employing repeated daily BDCM dosing regimens have also
    yielded results demonstrating liver and kidney toxicity.
    Thornton-Manning et al. (1994) administered five consecutive daily
    BDCM doses to female F344 rats and female C57BL/6J mice by aqueous
    gavage and found that BDCM is both hepatotoxic and nephrotoxic to
    female rats (150-300 mg/kg of body weight per day), but only
    hepatotoxic to female mice (75-150 mg/kg of body weight per day).
    Hepatic cytochrome P450 activities were decreased in rats, but not in
    mice, in this study. Munson et al. (1982) administered BDCM (50, 125
    or 250 mg/kg of body weight per day) to male and female CD-1 mice by
    aqueous gavage for 14 days and reported evidence for hepatic and renal
    toxicity as well as effects on the humoral immune system.
    Nephrotoxicity, as reflected by significant elevations of blood urea
    nitrogen (BUN), occurred only at the highest dose in both males and
    females. Male mice appeared more sensitive than females to
    BDCM-induced hepatotoxicity; 2- to 3-fold elevations in ASAT and ALAT
    (although not significant, according to the authors' statistical

    analysis) occurred at the lowest dose only in males. Based on the
    degree of these elevations, BDCM was the most potent hepatotoxicant
    compared with chloroform, DBCM and bromoform, which were also tested
    in this study. Immunotoxic effects described in the study included
    decreases in both antibody-forming cells and haemagglutination titres
    at the 125 and 250 mg/kg of body weight per day doses, although a
    recent investigation found no effects of BDCM on immune function
    (French et al., 1999). Condie et al. (1983) conducted a similar 14-day
    comparative dosing study with THMs and male CD-1 mice, but used corn
    oil as the vehicle of administration for doses of BDCM of 37, 74 and
    147 mg/kg of body weight per day. Evidence of renal damage was
    observed at the mid and high doses, whereas liver toxicity occurred
    only at the high dose. A 14-day corn oil gavage study by NTP (1987)
    demonstrated the greater sensitivity of the B6C3F1 mouse to BDCM: all
    male mice that received 150 or 300 mg/kg of body weight per day died
    before the end of the study. Aida et al. (1992a) incorporated
    microencapsulated BDCM into the diet of Wistar rats for 1 month, and a
    LOAEL of 66 mg/kg of body weight per day and a NOAEL of 21 mg/kg of
    body weight per day were determined based on histopathological
    findings of hepatocellular vacuolization.

         In a 13-week corn oil gavage study, NTP (1987) administered BDCM
    doses of 0, 19, 38, 75, 150 or 300 mg/kg of body weight per day, 5
    days per week, to F344/N rats (10 per sex per dose). The highest dose
    was lethal to 50% of males and 20% of females, and body weight
    depression was observed at the two highest doses. BDCM-induced lesions
    were found only at 300 mg/kg of body weight per day; these included
    hepatic centrilobular degeneration in both sexes and renal tubular
    degeneration and necrosis in males. Additional findings included mild
    bile duct hyperplasia and atrophy of the thymus, spleen and lymph
    nodes in both sexes. B6C3F1 mice were also dosed with BDCM in this
    study, and doses of 50 mg/kg of body weight per day and below produced
    no compound-related effects. Degeneration and necrosis of the kidney
    were observed in male mice at 100 mg/kg of body weight per day,
    whereas centrilobular degeneration of the liver was noted in females
    at 200 and 400 mg/kg of body weight per day.

    3)   Chronic toxicity

         Moore et al. (1994) administered BDCM in drinking-water
    (containing 0.25% Emulphor) to male F344 rats and B6C3F1 mice for
    1 year and evaluated clinical indicators of kidney toxicity. Water
    containing BDCM concentrations of 0.08, 0.4 and 0.8 g/litre for rats
    and 0.06, 0.3 and 0.6 g/litre for mice resulted in average daily doses
    of 4.4, 21 and 39 mg/kg of body weight for rats and 5.6, 24 and 49
    mg/kg of body weight for mice. A urinary marker for renal proximal
    tubule damage,  N-acetyl--glucosaminidase, was elevated above
    controls in each dose group in rats and at the highest treatment level
    in mice. Significant increases in urinary protein, indicative of
    glomerular damage, were also noted in low- and mid-dose rats as well
    as high-dose mice.

         In an NTP (1987) study, BDCM was administered by corn oil gavage
    for 102 weeks, 5 days per week, to F344/N rats (50 per sex per dose)
    at doses of 0, 50 or 100 mg/kg of body weight per day and to B6C3F1
    mice at doses of 0, 25 or 50 mg/kg of body weight per day (50 males
    per group) and 0, 75 or 150 mg/kg of body weight per day (50 females
    per group). In male rats, compound-related non-neoplastic lesions
    included renal cytomegaly and tubular cell hyperplasia and hepatic
    necrosis and fatty metamorphosis. Kidney tubule cell hyperplasia was
    also observed in female rats, as well as eosinophilic cytoplasmic
    change, clear cell change, focal cellular change and fatty
    metamorphosis of the liver. Histopathological changes were noted at
    both doses in rats. BDCM-induced non-neoplastic lesions in male mice
    included hepatic fatty metamorphosis, renal cytomegaly and follicular
    cell hyperplasia of the thyroid gland, all observed in both dose
    groups. In female mice, hyperplasia of the thyroid gland was observed
    at both doses.

         Microencapsulated BDCM was fed in the diet to Wistar rats for 24
    months, resulting in average daily doses of 6, 26 or 138 mg/kg of body
    weight for males and 8, 32 or 168 mg/kg of body weight for females
    (Aida et al., 1992b). Relative liver weight was increased in both
    sexes of all dose groups, as was relative kidney weight in the
    high-dose group. BDCM induced hepatic fatty degeneration and granuloma
    in all dose groups and cholangiofibrosis in the high-dose groups.
    Therefore, this study identified a LOAEL for chronic liver toxicity of
    6 mg/kg of body weight per day. 

    4.1.2.2  Reproductive and developmental toxicity

         Klinefelter et al. (1995) studied the potential of BDCM to alter
    male reproductive function in F344 rats. BDCM was consumed in the
    drinking-water for 52 weeks, resulting in average dose rates of 22 and
    39 mg/kg of body weight per day. No gross lesions in the reproductive
    organs were revealed by histological examination, but exposure to the
    high BDCM dose significantly decreased the mean straight-line, average
    path and curvilinear velocities of sperm recovered from the cauda
    epididymis. These effects of BDCM on sperm motility occurred at a
    lower exposure level than was observed for other DBPs that compromised
    sperm motility.

         A teratological assessment of BDCM was conducted in
    Sprague-Dawley rats by administering the compound by gavage from day 6
    to day 15 of gestation (Ruddick et al., 1983). Doses of 50, 100 and
    200 mg of BDCM per kg of body weight per day did not produce any
    teratogenic effects or dose-related histopathological changes in
    either the dams or fetuses, but sternebra aberrations were observed
    with a dose-dependent incidence in all dose groups. The increased
    incidence of these variations appeared to be significant, but no
    statistical analysis of the data was performed. Maternal weight gain
    was depressed in the high-dose group, and maternal liver and kidney
    weights were increased. Narotsky et al. (1997) employed a similar
    experimental model to test BDCM in F344 rats using doses of 0, 25, 50
    or 75 mg/kg of body weight per day in aqueous or oil gavage vehicles.
    BDCM induced full-litter resorptions in the 50 and 75 mg/kg of body

    weight per day dose groups with either vehicle of administration. For
    dams receiving corn oil, full-litter resorptions were noted in 8% and
    83% of the litters at 50 and 75 mg/kg of body weight per day,
    respectively. With the aqueous vehicle, 17% and 21% of the litters
    were fully resorbed at 50 and 75 mg/kg of body weight per day,
    respectively. All vehicle control litters and litters from the group
    given 25 mg/kg of body weight per day survived the experimental
    period. BDCM had been shown to cause maternal toxicity at these doses
    in a previous study (Narotsky et al., 1992).

    4.1.2.3  Neurotoxicity

         Neurotoxicological findings for the brominated THMs are limited
    to various observations of anaesthesia associated with acute high-dose
    exposures and results from a behavioural study conducted by Balster &
    Borzelleca (1982). Adult male ICR mice were dosed by aqueous gavage
    for up to 90 days. Treatments of 1.2 or 11.6 mg/kg of body weight per
    day were without effect in various behavioural tests, and dosing for
    30 days with 100 mg/kg of body weight per day did not affect passive
    avoidance learning. Animals dosed with either 100 or 400 mg/kg of body
    weight per day for 60 days exhibited decreased response rates in an
    operant behaviour test; these effects were greatest early in the
    regimen, with no evidence of progressive deterioration.

    4.1.2.4  Toxicity in humans

         Clinical case findings resulting from human exposure to BDCM have
    not been reported. 

    4.1.2.5  Carcinogenicity and mutagenicity

         IARC has evaluated the carcinogenicity of BDCM and concluded that
    there is sufficient evidence for its carcinogenicity in experimental
    animals and inadequate evidence for its carcinogenicity in humans. On
    this basis, BDCM was assigned to Group 2B: the agent is possibly
    carcinogenic to humans (IARC, 1991, 1999).

         Among the four THMs commonly found in drinking-water, BDCM
    appears to be the most potent rodent carcinogen. BDCM caused cancer at
    lower doses and at more target sites than for any of the other THMs.
    In the NTP (1987) 2-year bioassay, a corn oil gavage study (50 animals
    per sex per group, dosed 5 days per week), compound-related tumours
    were found in the liver, kidneys and large intestine. Daily doses were
    0, 50 or 100 mg/kg of body weight (male and female rats), 0, 25 or 50
    mg/kg of body weight (male mice) and 0, 75 or 150 mg/kg of body weight
    (female mice). NTP (1987) concluded that there was clear evidence of
    carcinogenic activity for both sexes of F344 rats and B6C3F1 mice, as
    shown by increased incidences of tubular cell adenomas and
    adenocarcinomas in the kidney and adenocarcinomas and adenomatous
    polyps in the large intestine of male and female rats, increased
    incidences of tubular cell adenomas and adenocarcinomas in the kidney
    of male mice, and increased incidences of hepatocellular adenomas and
    carcinomas in female mice (Table 11).

        Table 11. Tumour frequencies in rats and mice exposed to bromodichloromethane 
              in corn oil for 2 yearsa

                                                                                  
    Animal/tissue/tumour               Tumour frequency at control, low and high 
                                          doses (mg/kg of body weight per day)
                                                                                  

    Male rat                           0              50             100

    Large intestineb
        Adenomatous polyp              0/50           3/49           33/50
        Adenocarcinoma                 0/50           11/49          38/50
        Combined                       0/50           13/49          45/50

    Kidneyb
        Tubular cell adenoma           0/50           1/49           3/50
        Tubular cell adenocarcinoma    0/50           0/49           10/50
        Combined                       0/50           1/49           13/50

    Large intestine and/or kidney      0/50           13/49          46/50
      combinedb

    Female rat                         0              50             100

    Large intestinec
        Adenomatous polyp              0/46           0/50           7/47
        Adenocarcinoma                 0/46           0/50           6/47
        Combined                       0/46           0/50           12/47

    Kidney
        Tubular cell adenoma           0/50           1/50           6/50
        Tubular cell adenocarcinoma    0/50           0/50           9/50
        Combined                       0/50           1/50           15/50

    Large intestine and/or kidney      0/46           1/50           24/48
      combinedd

    Male mouse                         0              25             100

    Kidneye
        Tubular cell adenoma           1/46           2/49           6/50
        Tubular cell adenocarcinoma    0/46           0/49           4/50
        Combined                       1/46           2/49           9/50

    Female mouse                       0              75             150

    Liver
        Hepatocellular adenoma         1/50           13/48          23/50
        Hepatocellular carcinoma       2/50           5/48           10/50
        Combined                       3/50           18/48          29/50
                                                                                  

    Table 11. (continued)

    a   Adapted from NTP (1987).
    b   One rat died at week 33 in the low-dose group and was eliminated from 
        the cancer risk calculation.
    c   Intestine not examined in four rats from the control group and three 
        rats from the high-dose group.
    d   One rat in the high-dose group was not examined for intestinal tumours 
        and kidney tumours.
    e   In the control group, two mice died during the first week, one mouse 
        died during week 9 and one escaped in week 79. One mouse in the 
        low-dose group died in the first week. All of these mice were 
        eliminated from the cancer risk calculations. 
    

         Aida et al. (1992b) maintained Slc:Wistar rats on diets
    containing microencapsulated BDCM for 24 months and examined the
    animals for neoplastic lesions. The only significant finding was a
    slight increase in the incidence of liver tumours in females receiving
    the high dose (168 mg/kg of body weight per day). These included
    cholangiocarcinomas and hepatocellular adenomas.

         To date, effects following chronic BDCM administration via
    drinking-water have not been described in the literature. However, two
    separate drinking-water studies are currently being conducted by the
    US Environmental Protection Agency (EPA) and the NTP.

         Although BDCM has given mixed results in bacterial assays for
    genotoxicity, the results have tended to be positive in tests
    employing closed systems to overcome the problem of the compound's
    volatility (IARC, 1991, 1999; Pegram et al., 1997). Pegram et al.
    (1997) tested the THMs using a TA1535 strain transfected with rat GST
    T1-1 and found that the mutagenicity of the brominated THMs, but not
    chloroform, was greatly enhanced by the expression of the transferase.
    BDCM was mutagenic in this assay at medium concentrations below 0.1
    mmol/litre. Mutation spectra of the brominated THMs at the  hisG46 
    allele were characterized by DeMarini et al. (1997) using revertants
    induced in the GST-transfected strain. The overwhelming majority
    (96-100%) of the mutations induced by the brominated THMs were GC to
    AT transitions, and 87-100% of these were at the second position of
    the CCC/GGG target. BDCM produced primary DNA damage in the SOS
    chromotest  (Escherichia coli PQ37), but was negative in the Ames
    fluctuation test with  Salmonella typhimurium TA100 (Le Curieux et
    al., 1995). A mixture of BDCM and benzo [a]pyrene was tested in an
    Ames mutagenicity test with  S. typhimurium strains TA98 and TA100
    plus S9 (Kevekordes et al., 1998). BDCM in combination with
    benzo [a]pyrene caused a 25% increase in revertants in both strains
    compared with benzo [a]pyrene alone.

         BDCM was also positive in the majority of  in vitro genotoxicity
    tests employing eukaryotic systems, but the responses with and without
    an exogenous metabolizing system are less consistent. This may be due
    to the fact that the reactive intermediates suspected to be involved

    in THM mutagenicity must be generated within the target cells (Thier
    et al., 1993; Pegram et al., 1997). Moreover, extensive metabolism of
    the THMs by the supplemental S9 outside of the cells would greatly
    diminish the intracellular dose. Many of the positive studies are for
    the induction of sister chromatid exchange (SCE) (IARC, 1991, 1999).
    Morimoto & Koizumi (1983) found that BDCM induced SCEs in human
    lymphocytes  in vitro in the absence of S9 activation at
    concentrations greater than or equal to 0.4 mmol/litre, and Fujie et
    al. (1993) reported increased SCEs in rat erythroblastic leukaemia
    cells under similar conditions. Metabolically activated BDCM also
    increased SCEs  in vitro in human lymphocytes (at 1 mmol/litre) and
    in rat hepatocytes (at 100 mmol/litre) (Sobti, 1984).

          In vivo, doses of 50 mg of BDCM per kg of body weight and above
    produced SCEs in male CR/SJ mice (Morimoto & Koizumi, 1983). BDCM was
    negative in  in vivo clastogenicity tests (micronucleus formation) in
    mice and rats (Ishidate et al., 1982; Hayashi et al., 1988). Fujie et
    al. (1990) reported that BDCM induced bone marrow chromosomal
    aberrations (primarily chromatid and chromosome breaks) in Long-Evans
    rats following oral or intraperitoneal dosing at doses as low as 16.4
    mg/kg of body weight. Potter et al. (1996) found that BDCM did not
    induce DNA strand breaks in the kidney of male F344 rats following
    seven daily doses of 1.5 mmol/kg of body weight (246 mg/kg of body
    weight). Stocker et al. (1997) studied the effect of gastric
    intubation of aqueous solutions of BDCM on unscheduled DNA synthesis
    (UDS) in the liver of male rats. BDCM did not cause UDS in hepatocytes
    isolated after administration of single doses of 135 or 450 mg/kg of
    body weight. The  in vivo mutagenicity studies of BDCM and the other
    brominated THMs are summarized in Table 12.

         In comparison with other chemicals known to produce mutations via
    direct DNA reactivity such as aflatoxin B1 and ethylene dibromide,
    BDCM is a relatively weak mutagen.

    4.1.2.6  Comparative pharmacokinetics and metabolism

         Bromine substitution would be expected to confer greater
    lipophilicity on the brominated THMs compared with chloroform, which
    would affect tissue solubility and other factors that can influence
    pharmacokinetics. Because metabolism of each THM is qualitatively
    similar (with one known exception), this section addresses key
    features of the metabolism of all four THMs.

         The absorption, distribution and elimination of BDCM have been
    studied in rats and mice, and more recent work has led to the
    development of a physiologically based pharmacokinetic (PBPK) model
    for BDCM in rats. Mink et al. (1986) compared the pharmacokinetics of
    orally administered 14C-BDCM in male B6C3F1 mice and Sprague-Dawley
    rats. The animals received single doses of 100 mg/kg of body weight
    (rats) or 150 mg/kg of body weight (mice) in corn oil by gavage, and
    tissue levels of radioactivity were determined after 8 h. Absorption
    of BDCM appeared to be rapid and fairly complete, as would be expected
    for small halocarbons. This was especially true in the mouse, where
    93% of the dose was recovered within 8 h as carbon dioxide (81%), as


        Table 12. Dose information for selected  in vivo mutagenicity studies of brominated trihalomethanes

                                                                                                                                           
    End-point                      Assay system                        Dosea                        Result      Reference
                                                                                                                                           

    Sister chromatid exchange      Male CR/SJ mice, gavage, 4 days     50 mg BDCM/kg bw per day     positive    Morimoto & Koizumi (1983)

    Sister chromatid exchange      Male CR/SJ mice, gavage, 4 days     25 mg DBCM/kg bw per day     positive    Morimoto & Koizumi (1983)

    Sister chromatid exchange      Male CR/SJ mice, gavage, 4 days     25 mg bromoform/kg bw        positive    Morimoto & Koizumi (1983)
                                                                       per day

    Sister chromatid exchange      B6C3F1 mice, i.p.b                  200 mg bromoform/kg bw       positive    NTP (1989a)

    Micronucleus formation         ddY mice, MS mice, Wistar rats,     500 mg BDCM/kg bw per day    negative    Ishidate et al. (1982)
                                   i.p. in olive oil

    Micronucleus formation         ddY mice, MS mice, Wistar rats,     500 mg DBCM/kg bw per day    negative    Ishidate et al. (1982)
                                   i.p. in olive oil

    Micronucleus formation         ddY mice, MS mice, Wistar rats,     500 mg bromoform/kg bw       negative    Ishidate et al. (1982)
                                   i.p. in olive oil                   per day

    Micronucleus formation         ddY mice, i.p., single dose in      500 mg BDCM/kg bw            negative    Hayashi et al. (1988)
                                   corn oil

    Micronucleus formation         ddY mice, i.p., single dose in      1000 mg DBCM/kg bw           negative    Hayashi et al. (1988)
                                   corn oil

    Micronucleus formation         ddY mice, i.p., single dose in      1400 mg bromoform/kg bw      negative    Hayashi et al. (1988)
                                   corn oil

    Chromosomal aberrations        Long-Evans rats, bone marrow,       16.4 mg BDCM/kg bw           positive    Fujie et al. (1990)
                                   i.p., single dose

    Chromosomal aberrations        Long-Evans rats, bone marrow,       20.8 mg DBCM/kg bw           positive    Fujie et al. (1990)
                                   i.p., single dose
                                                                                                                                           

    Table 12. (continued)

                                                                                                                                           
    End-point                      Assay system                        Dosea                        Result      Reference
                                                                                                                                           

    Chromosomal aberrations        Long-Evans rats, bone marrow,       25.3 mg bromoform/kg bw      positive    Fujie et al. (1990)
                                   i.p., single dose

    Chromosomal aberrations        Long-Evans rats, bone marrow        253 mg bromoform/kg bw       positive    Fujie et al. (1990)
                                                                       per day

    Unscheduled DNA synthesis      Rat liver, gavage                   450 mg BDCM/kg bw per        negative    Stocker et al. (1997)
                                                                       day

    Unscheduled DNA synthesis      Rat liver, gavage                   2000 mg DBCM/kg bw per       negative    Stocker et al. (1997)
                                                                       day

    Unscheduled DNA synthesis      Rat liver, gavage                   1080 mg bromoform/kg bw      negative    Stocker et al. (1997)
                                                                       per day

    Micronucleus formation         Mouse, bone marrow, gavage,         1000 mg bromoform/kg bw      negative    Stocker et al. (1997)
                                   single dose

    DNA strand break               Male F344 rats, kidney, gavage,     1.5 mmol BDCM/kg bw per      negative    Potter et al. (1996)
                                   7 days                              day

    DNA strand break               Male F344 rats, kidney, gavage,     1.5 mmol DBCM/kg bw per      negative    Potter et al. (1996)
                                   7 days                              day

    DNA strand break               Male F344 rats, kidney, gavage,     1.5 mmol bromoform/kg bw     negative    Potter et al. (1996)
                                   7 days                              per day

    Sex-linked recessive           Drosophila                          1000 ppm solution            positive    NTP (1989a)
    mutation
                                                                                                                                           

    a  Doses listed are the lowest at which an effect was observed or, in the case of negative results, the highest dose tested. 
       bw = body weight.
    b  Intraperitoneal.
    

    expired volatile organics assumed to be unmetabolized parent compound
    (7.2%), in urine (2.2%) or in organs (3.2%). Much more of the 14C
    dose was expired as the assumed parent compound by the rat (42%) than
    by the mouse, but total recovery after 8 h was less (63%) because of
    lower conversion to carbon dioxide (14%). The liver, stomach and
    kidneys were the organs with the highest residual radioactivity
    levels. The authors estimated BDCM half-lives of 1.5 and 2.5 h in the
    rat and mouse, respectively.

         Mathews et al. (1990) studied the disposition of 14C-BDCM in
    male F344 rats after single oral (corn oil gavage) doses of 1, 10, 32
    or 100 mg/kg of body weight and 10-day repeat oral dosing of 10 or
    100 mg/kg of body weight per day. The doses of BDCM were well absorbed
    from the gastrointestinal tract, as demonstrated by 24-h recoveries
    exceeding 90% in non-faecal excreta samples and tissues. Persistence
    of radiolabelled residues in tissues after 24 h was low (3-4% of
    dose), with the most marked accumulation in the liver (1-3% of dose).
    The kidneys, particularly cortical regions, also contained significant
    concentrations of radiolabelled residues. Approximately 3-6% of the
    dose was eliminated as volatile organics in the breath (primarily the
    parent compound, presumably), much less than in Sprague-Dawley rats
    (Mink et al., 1986). Urinary and faecal elimination were low at all
    dose levels, accounting for 4% and 1-3% of the administered doses,
    respectively. Repeated doses had no effect on the tissue distribution
    of BDCM, and significant bioaccumulation was not observed (0.9-1.1%
    total retention of the label).

          Lilly et al. (1998) examined absorption and tissue dosimetry of
    BDCM in male F344 rats after doses of 50 or 100 mg/kg of body weight
    were given orally using either an aqueous emulsion or corn oil as the
    vehicle of administration. After delivery in the aqueous vehicle,
    concentrations of BDCM in venous blood peaked at about 6 min, with
    maximum concentration ( Cmax) values of 16 and 26 mg/litre for the
    low and high doses. With corn oil dosing,  Cmax occurred at 15-30
    min, with lower peak blood levels attained (5 and 9 mg/litre). The
    time required for blood concentrations to decline to half- Cmax was
    about 1 h with the 50 mg/kg of body weight dose and 1.5 h with the 100
    mg/kg of body weight dose. Tissue partition coefficient determinations
    confirmed the anticipated effect of bromine substitution on THM tissue
    solubility. BDCM partition coefficients for fat and liver were 526 and
    30.6 (Lilly et al., 1997b), compared with 203 and 21.1 for chloroform
    (Corley et al., 1990). Lilly et al. (1998) found slightly higher
    maximum concentrations of BDCM in the liver and kidneys after aqueous
    administration compared with corn oil delivery. With the 100 mg/kg of
    body weight aqueous dose, hepatic and renal levels peaked at about 15
    mg/litre at 5 min after dosing in the liver and at 5-30 min in the
    kidneys. At 6 h after dosing, concentrations of BDCM in the liver and
    kidneys were less than 1 mg/litre. More of the parent compound was
    eliminated unmetabolized via exhaled breath after aqueous dosing
    (8.9%, low dose; 13.2%, high dose) than after corn oil gavage (5.3%,
    low dose; 5.8%, high dose).

         The elimination kinetics of BDCM have been studied in humans who
    had swum in chlorinated pools (Lindstrom et al., 1997; Pleil &
    Lindstrom, 1997). BDCM half-lives of 0.45-0.63 min for blood were
    estimated using breath elimination data.

         The deleterious effects of the THMs result from reactive
    metabolites generated by biotransformation. Two isoenzyme groups of
    cytochrome P450, CYP2E1 and CYP2B1/2, as well as a theta-class GST,
    have been implicated in the metabolism of BDCM to toxic species in
    rats (Thornton-Manning et al., 1993; Pegram et al., 1997). No specific
    information is available regarding human metabolism of brominated
    THMs. In rats, P450-mediated metabolism of BDCM (Figure 1) is believed
    to proceed by the same two pathways established for chloroform:
    oxidation with phosgene the proposed active metabolite, and reduction
    with the dichloromethyl free radical proposed as the reactive product
    (Tomasi et al., 1985; IPCS, 1994; Gao et al., 1996). Most
    investigations of THM metabolism and reaction mechanisms have focused
    on chloroform, and a detailed review of chloroform biotransformation
    has been published recently (IPCS, 1994). Although the qualitative
    aspects of the cytochrome P450-mediated metabolism of brominated THMs
    are similar to those for chloroform, numerous studies have
    demonstrated that brominated THMs are metabolized to a greater extent
    and at faster rates than chloroform. There is evidence to suggest that
    both CYP2E1 and CYP2B1/2 can catalyse the oxidative pathway and that
    CYP2B1/2 catalyses the reductive metabolism of haloforms, but it has
    also been postulated that either isoform can catalyse both routes
    (Tomasi et al., 1985; Testai et al., 1996). CYP2E1 is clearly involved
    in the hepatotoxicity induced by BDCM in rats, but its role in the
    nephrotoxic response is less certain (Thornton-Manning et al., 1993).
    Because CYP2E1 is highly conserved across mammalian species, it seems
    likely that this isoform metabolizes BDCM in humans, although this has
    not yet been demonstrated. CYP2B1/2 are not expressed in humans, and
    there is no direct analogue for their catalytic activity; however,
    based on substrate similarities with CYP2B1/2, human forms CYP2A6,
    CYP2D6 and CYP3A4 appear to be possibilities. 

         Gao et al. (1996) demonstrated that GSH affords protection
    against the toxicity and macromolecular binding of BDCM, indicating
    that oxidative metabolism of BDCM, like that of chloroform, generates
    phosgene, which can then react with GSH (Stevens & Anders, 1981).
     In vitro binding of a BDCM-derived intermediate to microsomal
    protein under aerobic conditions and the prevention of this binding by
    GSH supplementation provide further evidence for the production of
    phosgene from BDCM (Gao et al., 1996). The reaction of GSH with
    phosgene forms  S-(chlorocarbonyl)-GSH, which may react with a second
    GSH molecule to produce diglutathionyl dithiocarbonate (Pohl et al.,
    1981) or to give glutathione disulfide and carbon monoxide as minor
    products (Stevens & Anders, 1981). Anders et al. (1978) and Mathews et
    al. (1990) reported that carbon monoxide is a product of BDCM
    metabolism. The most likely outcome for phosgene is hydrolysis to
    carbon dioxide and hydrogen chloride (Brown et al., 1974), and, in
    fact, 70-80% of 14C-BDCM doses administered to F344 rats or B6C3F1
    mice appeared as expired 14C-labelled carbon dioxide (Mink et al.,
    1986; Mathews et al., 1990), which also shows the predominance of

    FIGURE 1

    oxidative biotransformation as a metabolic route in these animals.
    Sprague-Dawley rats were much less efficient metabolizers of THMs,
    disposing of only 14% of a BDCM dose as carbon dioxide. In both rats
    and mice, BDCM was more extensively metabolized to carbon dioxide than
    was chloroform or bromoform (Mink et al., 1986).  In vitro binding
    assays with rat hepatic microsomes have also shown that BDCM has a
    greater capacity than chloroform to be metabolized to intermediates
    (presumably phosgene) that bind protein under aerobic conditions (Gao
    & Pegram, 1992; Bull et al., 1995). Similar tests with kidney
    microsomes have shown that GSH is much less effective in preventing
    renal protein binding than in preventing liver protein binding,
    suggesting that a significant portion of this binding in the kidney
    may have resulted from generation of a reactive intermediate via a
    different pathway (Gao et al., 1996).

         Cytochrome P450-mediated reductive dehalogenation of BDCM to form
    a dichloromethyl radical has been demonstrated  in vivo in
    phenobarbital-treated rats using an electron spin resonance (ESR)
    spin-trapping technique (Tomasi et al., 1985). These authors reported
    that more free radical was derived from BDCM than from chloroform and
    that more radical was trapped from bromoform than from BDCM. Waller &
    McKinney (1993) conducted a theoretical investigation into the
    potential of halogenated methanes to undergo reductive metabolism
    using density-functional theory-based computational chemistry. The
    estimated reductive potentials for the THMs were in agreement with the
    reaction order described in the ESR study. The dichloromethyl radical
    reacts preferentially with the fatty acid skeleton of phospholipids to
    give covalently bound adducts (De Biasi et al., 1992).  In vitro 
    binding of metabolically activated THMs to hepatic microsomal lipids,
    presumably by the radical, has also been investigated, and lipid
    binding by BDCM was found to exceed that of chloroform by more than
    300% (Gao & Pegram, 1992; Gao et al., 1996). Free radical generation
    by this pathway may explain the loss of CYP2B1/2 in rats treated with
    BDCM (Thornton-Manning et al., 1994). Finally, carbon monoxide has
    also been postulated to be a product of the reductive pathway (Wolf et
    al., 1977).

         Although the reactions of GSH with phosgene are protective
    against THM-induced hepatic and renal toxicity, direct conjugation of
    the brominated THMs to GSH may lead to genotoxicity. A GST-mediated
    mutagenic pathway of brominated THM metabolism has recently been
    identified using a  Salmonella strain transfected with rat GST T1-1
    (Pegram et al., 1997). Base substitution revertants were produced in
    this strain by BDCM (at medium concentrations of less than 0.1
    mmol/litre) and the other brominated THMs, but not by chloroform (De
    Marini et al., 1997; Pegram et al., 1997). The propensity for GSH
    conjugation may therefore explain the different results noted in
    mutagenicity tests with chloroform and the brominated THMs. Currently,
    it is not known if GSH conjugation leading to genotoxicity occurs in
    mammalian cells. However, it is likely that the analogous human GST
    T1-1 will also activate the brominated THMs, because similar substrate
    specificities have been demonstrated for the rat and human GST
    isoforms (Thier et al., 1993, 1996). Human GST T1-1 is expressed

    polymorphically and could therefore be a critical determinant of
    susceptibility to the genotoxicity of the brominated THMs.

         A PBPK model has recently been developed to describe the
    absorption, distribution, tissue uptake and dosimetry, metabolism and
    elimination of BDCM in rats (Lilly et al., 1997b, 1998). Metabolism
    was characterized directly by measuring production of bromide ion,
    which is liberated by all known biotransformation reactions of BDCM,
    and indirectly by gas uptake techniques. Determinations of plasma
    bromide concentrations after constant-concentration inhalation
    exposures of rats to BDCM provided evidence for metabolic saturation
    at concentrations of 1340 mg/m3 (200 ppm) and greater. Total
     in vivo metabolism of BDCM was accurately described by the model as
    a saturable process using metabolic rate constant values of 12.8 mg/h
    for maximum rate of metabolism ( Vmax) and 0.5 mg/litre for  Km.
    This compares with a  Vmax value for chloroform of 6.8 mg/h (Corley
    et al., 1990), providing additional evidence for more rapid metabolism
    and greater generation of reactive intermediates from BDCM than from
    chloroform. Production of bromide from BDCM following treatment with
    an inhibitor of CYP2E1,  trans-dichloroethylene, increased the
    apparent  Km from 0.5 to 225 mg/litre, further demonstrating that
    CYP2E1 is a major isoform involved in BDCM metabolism. The metabolism
    model, derived from inhalation exposure data, was subsequently linked
    to a multicompartment gastrointestinal tract submodel using estimates
    of oral absorption rate constants determined by fitting blood and
    exhaled breath chamber concentration-time curves obtained after gavage
    of rats with BDCM. This model accurately predicted tissue dosimetry
    and plasma bromide ion concentrations following oral exposure to BDCM
    and can be utilized in estimating rates of formation of reactive
    intermediates in target tissues.

    4.1.2.7  Mode of action

         As stated above, the metabolism of the THMs is believed to be a
    prerequisite for the toxicity and carcinogenicity associated with
    exposure to these DBPs. The primary target tissues for the THMs are
    active sites of their metabolism, and treatments that increase or
    decrease biotransformation also tend to cause parallel increases or
    decreases in the toxicity induced by the THMs (Thornton-Manning et
    al., 1993; US EPA, 1994b). The reactive intermediates generated from
    the three brominated THM metabolic pathways react with macromolecules
    to elicit both cytotoxic and genotoxic responses.

         The cytotoxicity of the brominated THMs observed in the liver and
    kidneys of exposed animals has been proposed to result from covalent
    adducts formed between cellular proteins and lipids and
    dihalocarbonyls or dihalomethyl free radicals. The adducts presumably
    impair the function of these molecules and cause cell injury. BDCM
    produced these adducts in  in vitro incubations with hepatic and
    renal microsomes to a significantly greater extent than did chloroform
    (Gao & Pegram, 1992; Gao et al., 1996). Induction of lipid
    peroxidation by free radical metabolites of reductive metabolism has
    been proposed as another mechanism underlying THM cytotoxicity. Each
    of the brominated THMs induced lipid peroxidation in rat liver

    microsomes  in vitro, which was maximal at low oxygen tensions (de
    Groot & Noll, 1989). 

         As described above (section 4.1.2.6), it appears that direct
    conjugation of BDCM and the other brominated THMs with GSH generates
    mutagenic intermediates (Pegram et al., 1997), but this process has
    not yet been demonstrated in mammals. However, the reaction was
    dependent on a transfected rat enzyme (GST T1-1), and the human GST
    T1-1 has been shown to catalyse GSH conjugation with the
    dihalomethanes (Thier et al., 1996). The GC to AT transitions observed
    in the  Salmonella strain expressing GST T1-1 indicate that DNA
    lesions on either guanine or cytosine occurred after exposure to
    brominated THMs (De Marini et al., 1997). Based on similar reactions
    by dihalomethanes, it can be proposed that  S-(1,1-dihalomethyl)-GSH
    is the product of GSH conjugation of brominated THMs, which then
    reacts directly with guanine. Dechert & Dekant (1996) found that
     S-(1-chloromethyl)-GSH reacts with deoxyguanosine to produce
     S-[1-( N2-deoxyguanosinyl)-methyl]-GSH, and a methylguanosine
    adduct was also formed from reactions of  S-(1-acetoxymethyl)-GSH
    with model nucleosides (Thier et al., 1993).  In vivo covalent DNA
    binding by 14C-BDCM has been observed in each of the cancer target
    tissues for BDCM in the rat, but the nucleoside adducts have not yet
    been identified (Bull et al., 1995).

         An interplay of direct mutagenicity with cytotoxic responses
    leading to regenerative hyperplasia may explain some, but not all, of
    the carcinogenic effects of the brominated THMs. For example, renal
    tubular cell hyperplasia coincident with tubular cell cancers was
    observed in rats gavaged with BDCM for 2 years (NTP, 1987), but
    necrosis and hyperplasia were not associated with the liver neoplasms
    induced by BDCM in mice (NTP, 1987). Melnick et al. (1998)
    recently revisited this issue and noted that high incidences of liver
    tumours were observed with BDCM and DBCM at doses that had little or
    no effect on hepatic regenerative hyperplasia. No evidence for
    cytotoxic responses in the intestine was noted in the NTP study with
    BDCM, but high incidences of intestinal carcinomas were reported (NTP,
    1987). Therefore, while cytotoxic effects of BDCM may potentiate
    tumorigenicity in certain rodent tissues at high dose levels, direct
    induction of mutations by BDCM metabolites may also play a
    carcinogenic role. The extent to which each of these processes
    contributed to the induction of tumours observed in chronic animal
    studies is at present unclear. Additional  in vivo studies are
    required to confirm the mechanism or mechanisms underlying brominated
    THM-induced carcinogenesis.

    4.1.3  Dibromochloromethane

    4.1.3.1  General toxicological properties and information on 
             dose-response in animals

    1)   Acute toxicity

         Acute oral LD50s of 800 and 1200 mg of DBCM per kg of body
    weight were reported by Bowman et al. (1978) for male and female ICR

    Swiss mice, respectively, whereas Chu et al. (1980) found LD50s of
    1186 and 848 mg/kg of body weight for male and female Sprague-Dawley
    rats, respectively. A DBCM dose of 500 mg/kg of body weight produced
    ataxia, sedation and anaesthesia in mice (Bowman et al., 1978). Hewitt
    et al. (1983) dosed male Sprague-Dawley rats with DBCM by corn oil
    gavage and found doses of 2450 mg/kg of body weight and above to be
    lethal. No clinical evidence for significant liver or kidney toxicity
    was found at sublethal doses.

         Induction of acute renal toxicity by DBCM was studied by Kroll et
    al. (1994a,b); a single intraperitoneal dose of 3 mmol/kg of body
    weight resulted in elevated BUN, reductions in glomerular filtration
    rate and renal concentrating ability, and interference with proximal
    tubular secretion and reabsorption.

    2)   Short-term toxicity

         Daily gavage of male and female CD-1 mice with DBCM in an aqueous
    vehicle for 14 days produced hepatotoxicity in both sexes at the
    highest dose of 250 mg/kg of body weight per day (Munson et al.,
    1982). Depressed immune function was also observed in both sexes at
    doses of 125 and 250 mg/kg of body weight per day, whereas the
    50 mg/kg of body weight per day dose was without effect. Corn oil
    gavage of DBCM to male CD-1 mice for 14 days (Condie et al., 1983) led
    to observations of kidney and liver toxicity at a lower dose
    (147 mg/kg of body weight per day) than had been observed with aqueous
    delivery (Munson et al., 1982). In another 14-day corn oil gavage
    study, NTP (1985) found that a dose of 500 mg/kg of body weight per
    day was lethal to B6C3F1 mice, and doses of 500 and 1000 mg/kg of
    body weight per day were lethal to F344/N rats. Dietary administration
    of microencapsulated DBCM to Wistar rats for 1 month caused liver cell
    vacuolization, with a LOAEL of 56 mg/kg of body weight per day and a
    NOAEL of 18 mg/kg of body weight per day (Aida et al., 1992a).

         DBCM-induced cardiotoxicity was reported in male Wistar rats
    after short-term exposure (4 weeks of daily dosing with 0.4 mmol/kg of
    body weight). Arrhythmogenic and negative chronotropic and dromotropic
    effects were observed, as well as extension of atrioventricular
    conduction times. Inhibitory actions of DBCM on calcium ion dynamics
    in isolated cardiac myocytes were also noted.

         In the NTP (1985) study, DBCM was administered by corn oil gavage
    to F344/N rats and B6C3F1 mice (10 per sex per dose) for 13 weeks (5
    days per week) at doses of 0, 15, 30, 60, 125 or 250 mg/kg of body
    weight per day. The highest dose was lethal to 90% of the rats,
    producing severe lesions and necrosis in kidney, liver and salivary
    glands. Hepatocellular vacuolization indicative of fatty changes was
    observed in male rats at doses of 60 mg/kg of body weight per day and
    higher. In the mice, no DBCM-related effects were reported at doses of
    125 mg/kg of body weight per day or lower. At the highest dose, fatty
    liver and toxic nephropathy were noted in males, but not in females. A
    NOAEL of 30 mg/kg of body weight per day can be derived from this
    study.

         A 90-day corn oil gavage study was conducted using Sprague-Dawley
    rats and doses of 0, 50, 100 or 200 mg/kg of body weight per day
    (Daniel et al., 1990b). Body weight gain was significantly depressed
    in the high-dose groups to less than 50% and 70% of the controls in
    males and females, respectively. Observations of liver damage included
    elevated ALAT in mid- and high-dose males, centrilobular lipidosis
    (vacuolization) in males at all doses and high-dose females, and
    centrilobular necrosis in high-dose males and females. Kidney proximal
    tubule cell degeneration was induced by DBCM in all high-dose rats and
    to a lesser extent at 100 mg/kg of body weight per day in males and at
    both 50 and 100 mg/kg of body weight per day in females.

    3)   Chronic toxicity

         The chronic oral toxicity of DBCM was studied by NTP (1985) in
    F344/N rats and B6C3F1 mice using corn oil gavage (5 days per week
    for 104 weeks) and doses of 0, 40 or 80 mg/kg of body weight per day
    for rats and 0, 50 or 100 mg/kg of body weight per day for mice. Liver
    lesions, including fat accumulation, cytoplasmic changes and altered
    basophilic staining, were observed in male and female rats at both
    dose levels. The low-dose male mice in this study were lost as a
    result of an overdosing accident. Compound-related hepatocytomegaly
    and hepatic focal necrosis were observed in high-dose male mice, and
    liver calcification (high dose) and fatty changes (both low and high
    doses) were noted in female mice. Renal toxicity (nephrosis) was also
    observed in male mice and female rats. 

    4.1.3.2  Reproductive and developmental toxicity

         Borzelleca & Carchman (1982) conducted a two-generation
    reproductive study of DBCM in ICR Swiss mice. Male and female mice at
    9 weeks of age were maintained on drinking-water containing 0, 0.1,
    1.0 or 4.0 mg of DBCM per ml, leading to average doses of 0, 17, 171
    or 685 mg/kg of body weight per day. Fertility and gestational index
    were reduced in the high-dose group for the F1 generations. Only
    fertility was decreased (high-dose) in the F2 generation. At the mid
    and high doses in both generations, litter size and the viability
    index were decreased. Other effects included decreased lactation index
    and reduced postnatal body weight. No dominant lethal or teratogenic
    effects were observed in the F1 or F2 generations.

         In a developmental study in rats conducted by Ruddick et al.
    (1983), gavage of DBCM (0, 50, 100, or 200 mg/kg of body weight per
    day) on gestational days 6-15 caused a depression of maternal weight
    gain, but no fetal malformations.

    4.1.3.3  Neurotoxicity

         DBCM was tested for behavioural effects in male ICR mice by
    Balster & Borzelleca (1982), who found no effect of treatments up to
    100 mg/kg of body weight per day for 60 days. Dosing for 60 days with
    400 mg/kg of body weight per day produced decreased response rates in
    an operant behaviour test. Korz & Gattermann (1997) observed
    DBCM-induced behavioural alterations in male golden hamsters exposed

    either for 14 days to a dose of 5 mg/kg of body weight per day or
    acutely to a single dose of 50 mg/kg of body weight. At the low dose,
    subchronic treatment caused reduced aggressive behaviour during social
    confrontation on day 14 as compared with vehicle-dosed controls.
    Following the acute dose, increased locomotor activity on days 3-6 and
    decreased wheel running on days 6-9 were observed, but no effects were
    noted after day 9.

    4.1.3.4  Toxicity in humans

         Clinical case findings resulting from human exposure to DBCM have
    not been reported. 

    4.1.3.5  Carcinogenicity and mutagenicity

         IARC has evaluated the carcinogenicity of DBCM and concluded that
    there is inadequate evidence for its carcinogenicity in humans and
    limited evidence for its carcinogenicity in experimental animals. The
    compound was assigned to Group 3: DBCM is not classifiable as to its
    carcinogenicity to humans (IARC 1991, 1999).

         In a 104-week corn oil gavage study, DBCM was not carcinogenic in
    F344 rats (50 per sex per dose) at doses of 0, 40 or 80 mg/kg of body
    weight given 5 days per week (NTP, 1985). In female B6C3F1 mice,
    however, DBCM significantly increased the incidence of hepatocellular
    adenomas and the combined incidences of hepatocellular adenomas and
    carcinomas at the high dose (100 mg/kg of body weight per day) (Table
    13). The incidence of hepatocellular carcinomas was significantly
    increased in male mice at the same dose. The low-dose (50 mg/kg of
    body weight per day) male mice were lost midway through the study as a
    result of an inadvertent overdose. NTP judged that these results
    provided equivocal evidence of DBCM carcinogenicity in male mice and
    some evidence of carcinogenicity in female mice.

         Based on  in vitro studies, mutagenic potency appears to
    increase with the degree of bromine substitution in the THMs. DBCM is
    mostly positive in tests employing closed systems to overcome the
    problem of volatility (IARC, 1991, 1999; Pegram et al., 1997). In the
    GST-transfected TA1535 strain (see section 4.1.1.3), DBCM is the most
    potent THM, inducing greater than 10-fold more revertants per plate
    than BDCM after exposure to THM vapour (3400 mg/m3 for DBCM; 2680
    mg/m3 for BDCM [400 ppm]) (Pegram et al., 1996; DeMarini et al.,
    1997). DBCM has given mostly positive results in eukaryotic test
    systems (Loveday et al., 1990; IARC, 1991, 1999; McGregor et al.,
    1991; Fujie et al., 1993), although there is less consistency in
    results between the different assays when considered with or without
    an exogenous metabolic system.

         Data from  in vivo studies are more equivocal. DBCM was positive
    for SCE and chromosomal aberrations in mouse bone marrow (Morimoto &
    Koizumi, 1983; Fujie et al., 1990) and in a newt micronucleus assay
    (Le Curieux et al., 1995), but was negative for micronuclei and UDS in
    the liver of rats (Ishidate et al., 1982; Hayashi et al., 1988;
    Stocker et al., 1997). Potter et al. (1996) found that DBCM did not

    Table 13. Frequencies of liver tumours in mice administered 
              dibromochloromethane in corn oil for 104 weeksa

                                                                          
    Treatment (mg/kg     Sex      Adenoma      Carcinoma     Adenoma or 
    of body weight                                           carcinoma 
    per day)                                                 (combined)
                                                                          

    Vehicle control      M        14/50        10/50         23/50
                         F        2/50         4/50          6/50

    50                   M        -b           -             -
                         F        4/49         6/49          10/49

    100                  M        10/50        19/50c        27/50d
                         F        11/50c       8/50          19/50e
                                                                          

    a  Adapted from NTP (1985).
    b  Male low-dose group was inadequate for statistical analysis.
    c   P < 0.05 relative to controls.
    d   P < 0.01 (life table analysis);  P = 0.065 (incidental tumour 
       tests) relative to controls.
    e   P < 0.01 relative to controls.



    induce DNA strand breaks in the kidneys of male F344 rats following
    seven daily doses of 1.5 mmol/kg of body weight.

         These studies are summarized in Table 12.

    4.1.3.6  Comparative pharmacokinetics and metabolism

         The pharmacokinetics of DBCM have been studied the least among
    the THMs. Mink et al. (1986) compared the absorption, distribution and
    excretion of DBCM with those of the other THMs. A dose of 100 mg/kg of
    body weight was administered orally in corn oil to male Sprague-Dawley
    rats by gavage, and 150 mg/kg of body weight was administered
    similarly to male B6C3F1 mice. The pattern of distribution and
    elimination of DBCM was very similar to that observed with BDCM: rats
    expired more of the dose than mice as a trapped organic component
    presumed to be the parent compound (48% vs. 12%) and expired less as
    carbon dioxide (18% vs. 72%) after 8 h. Total recovery from excised
    organs was 1.4% in rats and 5.0% in mice, whereas less than 2% of the
    dose was excreted in the urine in both species. Half-lives of DBCM in
    rats and mice were estimated to be 1.2 h and 2.5 h, respectively.

         The cytochrome P450-mediated metabolism of DBCM has not been
    directly investigated. Presumably, metabolism proceeds via the same
    routes of biotransformation as described for BDCM (section 4.1.2.6)
    and chloroform (IPCS, 1994). Oxidative metabolism of DBCM would be
    expected to yield a bromochlorocarbonyl rather than phosgene, and

    reductive dehalogenation would produce a bromochloromethyl radical.
    Anders et al. (1978) demonstrated  in vivo carbon monoxide production
    from DBCM in male Sprague-Dawley rats at a rate intermediate to those
    of BDCM and bromoform. DBCM was very reactive in the  Salmonella-GST
    mutagenicity assay, indicating that it has greater potential for GSH
    conjugation than BDCM (De Marini et al., 1997).

         Pankow et al. (1997) reported the metabolism of DBCM to bromide
    and carbon monoxide in rats after gavage of olive oil solutions
    (0.4-3.1 mmol/kg of body weight). The DBCM concentrations in blood and
    fat 6 h after the last of seven consecutive daily doses (0.8 mmol/kg
    of body weight) were lower than at 6 h after a single gavage of this
    dose, suggesting that more of the chemical was metabolized after the
    seventh dose than after the single dose. The seven-dose regimen also
    caused a 2-fold induction of a CYP2E1-specific activity. The
    involvement of CYP2E1, CYP2B1/2 and GSH in DBCM metabolism was also
    demonstrated.

    4.1.3.7  Mode of action

         Mechanistic issues for DBCM are similar to those addressed for
    BDCM (sections 4.1.2.6 and 4.1.2.7). The greater propensity for the
    metabolism of this compound and bromoform as compared with BDCM is
    difficult to reconcile with its lower carcinogenicity in the NTP
    (1985) bioassay. A possible explanation is less bioavailability
    resulting from the greater lipophilicity of this compound and the use
    of corn oil as the vehicle of administration. Greater lipophilicity
    and reactivity of this compound or its metabolites (i.e., the
    bromochlorocarbonyl metabolite) may also prevent it from reaching
    critical target sites. It is also of note that DBCM did not induce
    carcinomas of the large intestine in rats in the NTP studies using
    corn oil vehicle, whereas both BDCM and bromoform did induce these
    tumours. 

    4.1.4  Bromoform

    4.1.4.1  General toxicological properties and information on 
             dose-response in animals

    1)   Acute toxicity

         Among the brominated THMs, bromoform is the least potent as a
    lethal acute oral toxicant. The acute oral LD50s for bromoform were
    1400 and 1550 mg/kg of body weight in male and female mice,
    respectively (Bowman et al., 1978), and 1388 and 1147 mg/kg of body
    weight in male and female rats, respectively (Chu et al., 1980).
    Bromoform was also a less potent anaesthetic than BDCM and DBCM: a
    1000 mg/kg of body weight dose was required to produce this effect in
    mice. Intraperitoneal administration of bromoform in a corn oil
    vehicle at doses of 25-300 l/kg of body weight produced no
    significant elevations of serum enzymes indicative of liver damage
    (Agarwal & Mehendele, 1983).

         Bromoform was included in the acute renal toxicity studies of
    Kroll et al. (1994a,b) but was ranked as the least potent among the
    THMs in general disruption of renal function. A single intraperitoneal
    dose (3 mmol/kg of body weight) did, however, induce significant
    decreases in glomerular filtration rate, renal concentrating ability,
    and tubular secretion and reabsorption. Tubule function was affected
    as early as 8 h after dosing, whereas the other effects were observed
    at 24-48 h. 

    2)   Short-term toxicity

         Male and female CD-1 mice were gavaged daily with bromoform (50,
    125 or 250 mg/kg of body weight) in an aqueous vehicle for 14 days,
    leading to liver toxicity and a decrease in antibody-forming cells
    only at the highest dose (Munson et al., 1982). The magnitude of the
    liver effects was less than that observed with the other THMs, and BUN
    was not elevated by bromoform. Condie et al. (1983) noted both liver
    and kidney toxicity in CD-1 mice after dosing with 289 mg of bromoform
    per kg of body weight per day in corn oil for 14 days, whereas no
    significant effects were found at 72 and 145 mg/kg of body weight per
    day. NTP (1989a) also conducted a 14-day corn oil gavage study with
    bromoform and found that doses of 600 and 800 mg/kg of body weight per
    day were lethal to both sexes of F344/N rats. Mice given doses of 400
    mg/kg of body weight and higher had raised stomach nodules.
    Microencapsulated bromoform was added to the diet of Wistar rats for 1
    month, producing liver damage (Aida et al., 1992a). 

         Bromoform was given to F344/N rats (10 per sex per dose) and
    B6C3F1 mice (10 per sex per dose) by corn oil gavage for 13 weeks
    (5 days per week) at doses of 0, 25, 50, 100 or 200 mg/kg of body
    weight per day, with an added dose of 400 mg/kg of body weight per day
    for mice only (NTP, 1989a). The only significant finding, hepatic
    vacuolation, was observed in male rats, but not in female rats, at
    doses of 50 mg/kg of body weight per day and higher, and in male mice,
    but not in female mice, at doses of 200 and 400 mg/kg of body weight
    per day. The NOAEL in the rat study was 25 mg/kg of body weight per
    day, whereas that in the study in mice was 100 mg/kg of body weight
    per day.

    3)   Chronic toxicity

         Bromoform was administered in drinking-water (containing 0.25%
    Emulphor) to male F344 rats and B6C3F1 mice for 1 year, and clinical
    indicators of kidney toxicity were examined (Moore et al., 1994).
    Water containing bromoform concentrations of 0.12, 0.6 and 1.2 g/litre
    for rats and 0.08, 0.4, and 0.8 g/litre for mice resulted in average
    daily doses of 6.2, 29 or 57 mg/kg of body weight for rats and 8.3, 39
    or 73 mg/kg of body weight for mice. Several indicators of tubular and
    glomerular damage were elevated at each treatment level in mice, and
    mice appeared more susceptible to the nephrotoxic effects of bromoform
    than to those of BDCM (see section 4.1.2.1). As in mice, urinary
    protein was increased in all rat dose groups, but little evidence of
    loss of tubule function was observed in rats.

         Two-year investigations of bromoform toxicity were conducted by
    administering doses of 0, 100 or 200 mg/kg of body weight by corn oil
    gavage, 5 days per week for 103 weeks, to F344/N rats (50 per sex per
    dose) and female B6C3F1 mice (50 per dose) (NTP, 1989a). Male mice
    (50 per dose) received doses of 0, 50 or 100 mg/kg of body weight per
    day. Survival of high-dose male rats and both dose groups of female
    mice was significantly lower than that of the vehicle controls. Focal
    or diffuse fatty change of the liver was observed in a dose-dependent
    fashion in both sexes of rats, and active chronic hepatic inflammation
    was found in male and high-dose female rats. Minimal necrosis of the
    liver was noted in high-dose male rats. In mice, the incidence of
    fatty changes of the liver was increased in females, but not in males,
    at both dose levels. Follicular cell hyperplasia of the thyroid gland
    was observed in high-dose female mice. The different target organ
    (liver) in this study as compared with the drinking-water study of
    Moore et al. (1994), in which renal effects predominated, suggests
    that the vehicle and mode of administration can affect which tissues
    are affected by bromoform.

    4.1.4.2  Reproductive and developmental toxicity

         Ruddick et al. (1983) conducted a teratological investigation of
    bromoform in Sprague-Dawley rats (15 per group). Bromoform was
    administered by gavage at doses of 50, 100 or 200 mg/kg of body weight
    per day on gestation days 6-15. Evidence of a fetotoxic response was
    observed, but there were no fetal malformations. Interparietal
    deviations were, however, noted in the mid- and high-dose groups. NTP
    (1989b) utilized the same bromoform doses given by gavage for 105 days
    to 20 male-female pairs of Swiss CD-1 mice to examine effects on
    fertility and reproduction. There was no detectable effect of
    bromoform on fertility, litters per pair, live pups per litter,
    proportion of pups born alive, sex of live pups or pup body weights.
    Bromoform was also found to induce full-litter resorptions in pregnant
    F344 rats when administered orally on gestation days 6-15, but at
    higher doses (150 and 200 mg/kg of body weight per day) than those
    required to produce the same effect for BDCM (Narotsky et al., 1993).

    4.1.4.3  Neurotoxicity

         Bromoform was included in the mouse behavioural study of Balster
    & Borzelleca (1982). Doses of 9.2 mg/kg of body weight per day for up
    to 90 days had no effect on the outcome of several behavioural tests,
    and 100 mg/kg of body weight per day for 30 days did not deter passive
    avoidance learning. However, mice receiving either 100 or 400 mg/kg of
    body weight per day for 60 days exhibited decreased response rates in
    an operant behaviour test.

    4.1.4.4  Toxicity in humans

         Bromoform was used in the late 19th and early 20th centuries as a
    sedative for children with whooping cough. Patients were typically
    given doses of one drop (approximately 180 mg) 3-6 times per day
    (Burton-Fanning, 1901), which usually resulted in mild sedation in the
    children. A few rare instances of death or near-death were reported
    but were believed to be due to accidental overdoses (Dwelle, 1903).

    These clinical observations have been used to estimate a lethal dose
    for a 10- to 20-kg child to be about 300 mg/kg of body weight and an
    approximate minimal dose for sedation to be 50 mg/kg of body weight
    per day (US EPA, 1994b).

    4.1.4.5  Carcinogenicity and mutagenicity

         IARC has evaluated the carcinogenicity of bromoform and concluded
    that there is inadequate evidence for its carcinogenicity in humans
    and limited evidence for its carcinogenicity in experimental animals.
    The compound was assigned to Group 3: bromoform is not classifiable as
    to its carcinogenicity to humans (IARC, 1991, 1999).

         Two-year studies of bromoform carcinogenicity were conducted by
    administering doses of 0, 100 or 200 mg/kg of body weight in corn oil
    by gavage, 5 days per week, to groups of F344/N rats (50 per sex per
    group) and female B6C3F1 mice (50 per group) and doses of 0, 50 or
    100 mg/kg of body weight to male mice (NTP, 1989a). No neoplastic
    effects were associated with the exposure of mice to bromoform. In
    rats, however, intestinal carcinomas were induced by bromoform, as had
    also been observed with BDCM. The uncommon adenomatous polyps or
    adenocarcinomas (combined) of the large intestine (colon or rectum)
    were observed in 3 out of 50 high-dose male rats and in 8 out of 50
    high-dose female rats as compared with 0 out of 50 rats in the male
    controls and 0 out of 48 rats in the female controls (Table 14).
    Because these tumours are very rare in the rat, these findings were
    considered significant, and the NTP therefore concluded that there was
    clear evidence for carcinogenic activity in female rats and some
    evidence in male rats.

         Bromoform, in common with the other brominated THMs, is largely
    positive in bacterial assays of mutagenicity conducted in closed
    systems (Zeiger, 1990; IARC, 1991, 1999). In the GST-transfected
     Salmonella typhimurium TA1535 strain (see section 4.1.2.5),
    bromoform produced about 5 times more revertants than BDCM at
    comparable exposure levels (Pegram et al., 1997). As with BDCM, the
    mutations were almost exclusively GC to AT transitions (DeMarini et
    al., 1997). In prokaryotic tester strains, bromoform induced mutations
    without metabolic activation in  S. typhimurium strain TA100 (Simmon
    & Tardiff, 1978; Ishidate et al., 1982; NTP, 1989a; Le Curieux et al.,
    1995), with and without activation in TA98 (NTP, 1989a; Zeiger, 1990)
    and with microsomal activation in TA97 (NTP, 1989a). In eukaryotic
    test systems, bromoform is largely positive (IARC, 1991; Fujie et al.,
    1993). As with bacterial assays, bromoform appeared more potent than
    the other brominated THMs (Morimoto & Koizumi, 1983; Banerji &
    Fernandes, 1996). 

          In vivo studies (summarized in Table 12) have given
    contradictory results. Bromoform was positive and negative in
     Drosophila (Woodruff et al., 1985). It was positive in the newt
    micronucleus test (Le Curieux et al., 1995) and gave increased SCE and
    chromosomal aberrations in mouse and rat bone marrow cells (Morimoto &
    Koizumi, 1983; Fujie et al., 1990). It gave negative results in mouse
    bone marrow (Hayashi et al., 1988; Stocker et al., 1997), in the rat

    Table 14. Tumour frequencies in rats exposed to bromoform in corn oil 
              for 2 yearsa

                                                                          
    Animal/tissue/tumour                   Tumour frequency
                                                                          
                                Control      100 mg/kg of    200 mg/kg of 
                                             body weight     body weight 
                                             per day         per day
                                                                          

    Male rat

    Large intestine
        Adenocarcinoma          0/50         0/50            1/50
        Polyp (adenomatous)     0/50         0/50            2/50

    Female rat

    Large intestine
        Adenocarcinoma          0/48         0/50            2/50
        Polyp (adenomatous)     0/48         1/50            6/50
                                                                          

    a Adapted from NTP (1989a).



    liver UDS assay (Pereira et al., 1982; Stocker et al., 1997) and in
    the dominant lethal assay (Ishidate et al., 1982). In the studies
    carried out by the NTP (1989a), it was positive for micronuclei and
    SCE, but negative for chromosomal aberrations in mouse bone marrow.
    Potter et al. (1996) found that bromoform did not induce DNA strand
    breaks in the kidneys of male F344 rats following seven daily doses of
    1.5 mmol/kg of body weight.

    4.1.4.6  Comparative pharmacokinetics and metabolism

         14C-Bromoform pharmacokinetics were examined in Sprague-Dawley
    rats and B6C3F1 mice in a study that included all four of the common
    THMs that occur as DBPs (Mink et al., 1986). Recoveries of 14C-label
    8 h after gavage dosing were 79% in rats and 62% in mice. The
    distribution and elimination of bromoform resembled those of
    chloroform rather than those of the other brominated THMs. The
    percentage of 14C-label recovered from excised organs and tissues was
    2.1% in rats and 12.2% in mice. Tissue levels of 14C in mice were
    substantially greater than those observed for the other brominated
    THMs. Urinary excretion of label after 8 h (2.2-4.6%) was also greater
    for bromoform than for BDCM and DBCM. Bromoform (and organic
    metabolite) elimination via exhaled breath was greater than that for
    all other THMs in the rat (67%), but less than that for all other THMs
    in the mouse (6%). The estimated half-life of bromoform was 0.8 h in
    rats and 8 h in mice.

         The cytochrome P450-dependent metabolism of bromoform was studied
    by Anders et al. (1978), who found that bromoform was converted to
    carbon monoxide  in vivo in male Sprague-Dawley rats at a rate
    significantly greater than were the other three THMs. Evidence was
    presented suggesting that some of the carbon monoxide arose from the
    oxidative metabolism of bromoform to a dibromocarbonyl and subsequent
    reactions with GSH. It is possible that some of the carbon monoxide
    was also generated from reductive metabolism of bromoform, as
    indicated by  in vitro results of Wolf et al. (1977) from anaerobic
    incubations using hepatic preparations derived from
    phenobarbital-treated rats. Carbon monoxide was produced from
    anaerobic bromoform metabolism in this study at much greater levels
    than from chloroform metabolism. Among the four THMs tested by Mink et
    al. (1986), bromoform exhibited the least metabolism to carbon dioxide
    (4% in rats and 40% in mice in 8 h). Free radicals from the  in vivo 
    reductive metabolism of bromoform were detected by ESR spin-trapping
    after dosing phenobarbital-treated rats (Tomasi et al., 1985). More
    radicals were generated from bromoform  in vivo than from any other
    bromochlorinated THM, which is consistent with computational chemistry
    predictions that bromoform would have the greatest reductive potential
    of the THMs (Waller & McKinney, 1993). Knecht & Mason (1991) also
    detected the  in vivo production of radicals from bromoform in the
    bile of rats, but only after the induction of hypoxia. 

         Bromoform, like DBCM, has a much greater potential than BDCM to
    be conjugated by GSH to form a mutagenic intermediate. Base pair
    revertants were produced in a dose-dependent fashion in
    GST-transfected  Salmonella typhimurium strain TA1535 by bromoform
    and BDCM, but not by chloroform (Pegram et al., 1996, 1997; De Marini
    et al., 1997). At an exposure concentration of 33 100 mg/m3 for
    bromoform and 22 800 mg/m3 for BDCM (3200 ppm), GST-dependent
    revertants per plate induced by bromoform and BDCM averaged 373 and
    1935, respectively, compared with a control rate of 23 per plate
    (Pegram et al., 1996).

    4.1.4.7  Mode of action

         The basic mechanisms of action for bromoform are similar to those
    described for BDCM (section 4.1.2.7). Although bromoform seems to have
    a greater propensity for metabolism and is a more potent mutagen than
    BDCM, it appears to be a less potent toxicant and carcinogen based on
    the results of the NTP (1985, 1987) bioassays and numerous other
     in vivo studies of toxicity. As with DBCM, a possible explanation is
    less bioavailability resulting from the greater lipophilicity of this
    compound and the use of corn oil as the vehicle of administration.
    This concept may be supported by the occurrence of bromoform-induced
    tumours in the intestinal tract, but not in the liver or kidneys.
    Greater lipophilicity and reactivity of bromoform metabolites may also
    prevent it from reaching critical target sites. Moreover, when
    bromoform was injected intraperitoneally, its metabolism was greater
    than that of the other THMs (Anders et al., 1978; Tomasi et al.,
    1985); when administered by corn oil gavage, however, bromoform was
    the least metabolized THM (Mink et al., 1986). Data from studies of
    oral exposure to bromoform in aqueous solutions are therefore
    required.

    4.2  Haloacids

         Like the THMs, the haloacids produced in the chlorination of
    drinking-water consist of a series of chlorinated and brominated
    forms. To date, the chlorinated acetic acids have been more thoroughly
    characterized toxicologically than their brominated analogues. As
    discussed in earlier chapters, the dihaloacetates and trihaloacetates
    occur in significantly higher concentrations than the
    monohaloacetates. The present review will emphasize the di- and
    trihaloacetates as the dominant forms of the haloacids found in
    drinking-water and the ones for which extensive toxicological data
    have been developed. The probable existence of many longer-chain
    halogenated acids in chlorinated drinking-water may be surmised from
    studies of chlorinated humic acids (reviewed by Bull & Kopfler, 1991).
    Few of these compounds have been extensively studied toxicologically.
    However, brief reference will be made to those compounds that are
    known to share some of the effects of the HAAs on intermediary
    metabolism.

    4.2.1  Dichloroacetic acid (dichloroacetate)

         Before beginning discussions of available toxicological
    evaluations of DCA, it is important to point out that this compound
    exists in drinking-water as the salt, despite the fact that it is
    widely referred to as dichloroacetic acid. DCA has a p Ka of 1.48 at
    25C (IARC, 1995). As a consequence, it occurs almost exclusively in
    the ionized form at the pHs found in drinking-water (broadly speaking,
    a pH range of 5-10). Failure to recognize this has resulted in a
    number of studies that have employed the free acid in test systems. At
    low doses, the buffering capacity of the physiological system can
    neutralize acid, and the measured activity may in fact be
    representative. However, most of the experimentation has been
    conducted using doses ranging from 50 to 1000 mg/kg of body weight
     in vivo or in the mmol/litre range  in vitro. Therefore, the
    applicability of the results of such studies to estimating human risks
    will be uncertain because of the large pH artefacts that can be
    expected when administering these quantities of a strong acid.

         DCA has been shown to produce developmental, reproductive, neural
    and hepatic effects in experimental animals. In general, these effects
    occur when the compound has been administered at high dose rates, at
    which there is evidence to indicate that the metabolic clearance of
    DCA is substantially inhibited. This has important implications in
    attempting to associate these toxicities with the low doses that are
    obtained in drinking-water.

         Because it was being developed as a potential therapeutic agent,
    there is a toxicological literature that precedes DCA's discovery as a
    by-product of chlorination. The present review discusses these early,
    more general explorations of DCA's effects on intermediary metabolism
    before proceeding to descriptions of studies designed to more
    specifically study its toxicology.

    4.2.1.1  General toxicological properties and information on 
             dose-response in animals

    1)   Acute toxicity

         DCA is not very toxic when administered acutely to rodents.
    Woodard et al. (1941) reported LD50s of 4.5 and 5.5 g/kg of body
    weight in rats and mice, respectively, for DCA administered as the
    sodium salt. This is roughly in the same range as the LD50s for
    acetic acid. There is reason to believe that other species, most
    specifically the dog, may be more sensitive because of some
    repeated-dose experiments discussed below; however, no specific data
    seem to have been reported in the literature.

    2)   Short-term toxicity

         Katz et al. (1981) published the first substantive evaluation of
    DCA's subchronic toxicity in rats and dogs. DCA was administered by
    gavage at 0, 125, 500 or 2000 mg/kg of body weight per day to rats
    (10-15 per sex per group) and at 0, 50, 75 or 100 mg/kg of body weight
    per day to dogs (3-4 per sex per group) for 3 months. One of three
    female dogs died at 75 mg/kg of body weight per day, and one of four
    male dogs died at 100 mg/kg of body weight per day. The most overt
    toxicity in rats was hindlimb paralysis at the highest dose. Clinical
    chemistry indicated significant increases in total and direct
    bilirubin at 500 and 2000 mg/kg of body weight per day in rats, and
    relative liver weights were significantly increased at all doses. In
    dogs, an increase in the incidence of haemosiderin-laden Kupffer cells
    was noted at all dose rates. Histopathological changes were observed
    in the brain and testes of both species. In rats, oedematous brain
    lesions were seen at 60% incidence (primarily in the cerebrum, but
    also in the cerebellum) at the lowest dose and at 100% in the two
    higher doses in both sexes. Slight to moderate vacuolation of
    myelinated tracts was observed at all doses in both rats and dogs.
    Testicular germinal epithelial degeneration was observed in rats at
    doses of 500 mg/kg of body weight per day and above and at all doses
    in dogs, with severity increasing with dose. In dogs, a high incidence
    of ocular anomalies was observed, and lenticular opacities were found
    to be irreversible upon suspension of treatment with DCA. In both rats
    and dogs, glucose and lactate levels were suppressed in a dose-related
    manner.

         Bull et al. (1990) examined the effects of DCA on the liver of
    B6C3F1 mice administered 1 or 2 g of DCA per litre (approximately 170
    and 300 mg/kg of body weight per day) in their drinking-water, with
    exposure lasting up to 1 year. As had been at least alluded to in
    earlier studies, DCA was found to produce a severe hepatomegaly in
    mice at concentrations in drinking-water of 1 g/litre and above. The
    hepatomegaly could be largely accounted for by large increases in cell
    size (cytomegaly). In shorter-term experiments, it was determined that
    treatment with DCA produced only minor changes in the labelling index
    of hepatocytes using a pulse dose of tritiated thymidine (Sanchez &
    Bull, 1990). In general, increases in labelling indices were seen in
    areas of acinar necrosis in the liver. Hepatocytes from these mice

    stained very heavily for glycogen using periodic acid/Schiff's reagent
    (PAS). The accumulation of glycogen began to occur with as little as
    1-2 weeks of treatment (Sanchez & Bull, 1990), but became
    progressively more severe with time (Bull et al., 1990). Sanchez &
    Bull (1990) noted that these effects could not be replicated by
    exposing mice to the metabolites of DCA -- glycolate, glyoxylate or
    oxalate -- in the drinking-water.

         Davis (1990) investigated the effects of DCA and TCA treatments
    on male and female Sprague-Dawley rats for up to 14 days. The authors
    reported decreased plasma glucose and lactic acid concentrations in
    both plasma and liver. There were some inconsistencies in the
    reporting of the doses in this study. However, the author has provided
    the correct doses, which were 120 and 316 mg/kg of body weight per day
    (M.E. Davis, personal communication, 1996). This places the results in
    a context that is more consistent with those of other studies (e.g.,
    Katz et al., 1981).

         Mather et al. (1990) found increases in liver weight in rats
    treated with DCA at 5 g/litre (350 mg/kg of body weight per day) in
    their drinking-water for 90 days. Relative liver and kidney weights
    were increased at concentrations of 0.5 g/litre (35 mg/kg of body
    weight per day) and above. PAS staining of liver sections revealed
    accumulation of glycogen in severely swollen hepatocytes, which was
    quite marked with 5 g DCA/litre. The treatment caused small,
    statistically significant increases in alkaline phosphatase (AP) and
    ALAT in serum. However, these changes were too small to be of clinical
    importance. DCA was found to approximately double the activity of
    cyanide-insensitive acyl coenzyme A (CoA) activity in the liver at
    5 g/litre, indicating some induction of peroxisome synthesis.

         Cicmanec et al. (1991) examined the subchronic effects of DCA in
    dogs. DCA was administered at doses of 0, 12.5, 39.5 or 72 mg/kg of
    body weight per day for 90 days to groups of five males and five
    females. Liver weights were significantly increased in a dose-related
    manner, beginning with the lowest dose, and kidney weight was
    increased at the highest dose. This was accompanied by
    histopathological observation of vacuolar changes in the liver and
    haemosiderosis. The pancreas displayed evidence of chronic
    inflammation and acinar degeneration at the two highest doses.
    Testicular degeneration was observed in virtually all the male dogs
    administered DCA. This pathology was not observed in the control
    animals. Vacuolization of white myelinated tracts in the cerebrum or
    cerebellum was observed at all doses, and vacuolar changes were
    observed in the medulla and spinal cord of male dogs. Although the
    vacuolization was present in all dose groups, the authors described it
    as being mild. These authors failed to find evidence of the lenticular
    opacities that had been previously reported by Katz et al. (1981)
    under very similar treatment conditions. They also found fairly
    consistent decreases in erythrocyte counts and haemoglobin
    concentrations at 72 mg/kg of body weight per day in both male and
    female dogs. As found by others in rat studies, evidence of hindlimb
    paralysis was reported, but the effect was expressed only sporadically
    in dogs.

         A number of other studies are consistent with the pattern of
    toxicological effects produced by DCA described above. Included in
    this group are the study of Bhat et al. (1991), which adds
    observations of enlarged portal veins, deposition of collagen in the
    area of the portal triads and some similar lesions in the vasculature
    of the lung in rats. This study also noted atrophic testes and focal
    vacuolation and gliosis in the brain. Only a single dose level of 10.5
    g/litre of drinking-water (approximately 1100 mg/kg of body weight per
    day) was utilized in this 90-day study, a level that substantially
    exceeds concentrations reported to reduce water and food consumption
    in other studies (Bull et al., 1990).

         More recent studies have more closely examined the dose-response
    relationships involved in the accumulation of glycogen in the liver of
    male B6C3F1 mice treated with DCA in drinking-water at levels ranging
    from 0.1 to 3 g/litre (approximately 20-600 mg/kg of body weight per
    day) for up to 8 weeks (Kato-Weinstein et al., 1998). Significant
    increases in the glycogen content of the liver of mice were seen with
    concentrations as low as 0.5 g/litre (100 mg/kg of body weight per
    day) in their drinking-water, with a small, but insignificant,
    increase being observed at 0.2 g/litre (40 mg/kg of body weight per
    day). Glycogen concentrations in the liver reached maximum levels
    within 1 week of treatment with concentrations in drinking-water of
    1 g/litre and above. At this early stage, the glycogen that
    accumulates is subject to mobilization by fasting. With continued
    treatment, the glycogen that is deposited becomes increasingly
    resistant to mobilization until approximately 8 weeks of treatment,
    when the glycogen contents of the livers of DCA-treated mice are not
    different in fasted and non-fasted states.

         The enzymatic basis of hepatic glycogen accumulation remains
    unclear. DCA treatment has no effect on the total amount of either
    form of glycogen synthase in the liver (e.g.,
    glucose-6-phosphate-dependent vs. glucose-6-phosphate-independent
    activity). The proportion of glycogen synthase in the active form was
    significantly decreased in mice treated with DCA for as little as
    1 week. The amount of phosphorylase in the active form appeared
    unaltered by treatment. Such changes could indicate that a feedback
    inhibition may have developed on the synthesis of glycogen as a result
    of its accumulation in hepatocytes.

         Carter et al. (1995) examined the time course of DCA's effects in
    the liver of B6C3F1 mice at concentrations of 0, 0.5 or 5 g/litre in
    drinking-water for up to 30 days. As reported in prior studies, the
    high-dose group displayed severe liver hypertrophy. However, a
    smaller, but consistent, increase in liver weight became evident with
    as little as 10 days of treatment at 0.5 g/litre. Even at this
    relatively low dose, some hepatocytes appeared to have lost nuclei or
    possessed nuclei that had undergone some degree of karyolysis. These
    experiments also appeared to rule out cytotoxicity and reparative
    hyperplasia as consistent features of DCA's effects. The authors
    suggested that this was in apparent contrast to the earlier
    observations of Sanchez & Bull (1990). However, Sanchez & Bull (1990),
    in a 14-day study in mice, noted that these effects were closely

    associated with what appeared to be infarcted areas in the liver of
    mice treated with high doses of DCA rather than cytotoxicity. These
    infarcts are thought to be secondary to the severe swelling of
    hepatocytes that results from DCA treatment. This interpretation is
    supported by the apparent lack of cytotoxic effects of DCA in isolated
    hepatocytes of both mice and rats at concentrations in the mmol/litre
    range (Bruschi & Bull, 1993). Subsequently, these infarcted areas have
    been identified as acinar necrosis (ILSI, 1997), which occurs in a
    somewhat random fashion when high concentrations of DCA (> 2
    g/litre) are administered for prolonged periods of time (Stauber &
    Bull, 1997).

         It is important to note that the dog appears to be very sensitive
    to the effects of DCA on the liver; substantial increases in liver
    weights are observed at daily doses as low as 12.5 mg/kg of body
    weight per day for a 90-day period (Cicmanec et al., 1991). By
    comparison, the lowest effect level noted in mice is 0.5 g/litre of
    drinking-water, which approximates 70-100 mg/kg of body weight per day
    (Carter et al. 1995), and the lowest effect level noted in rats is 125
    mg/kg of body weight per day (Katz et al., 1981).

    4.2.1.2  Reproductive effects

         DCA produces testicular toxicity when administered at high doses
    in drinking-water. These effects were first noted in studies of the
    general toxicity of DCA (Katz et al., 1981) and were discussed above
    (section 4.2.1.1). Cicmanec et al. (1991) followed up on these
    original observations in dogs and detected degeneration of the
    testicular epithelium and syncytial giant cell formation at doses as
    low as 12.5 mg/kg of body weight.

         Toth et al. (1992) examined DCA's ability to modify male
    reproductive function in Long-Evans rats given 0, 31, 62 or 125 mg/kg
    of body weight per day by gavage for 10 weeks. Reduced weights of
    accessory organs (epididymis, cauda epididymis and preputial gland)
    were observed at doses as low as 31 mg/kg of body weight per day.
    Epididymal sperm counts were found to be depressed and sperm
    morphology was increasingly abnormal at doses of 62 mg/kg of body
    weight per day and above. These latter effects were accompanied by
    changes in sperm motion. Fertility was tested in overnight matings and
    was found to be depressed only at the highest dose evaluated,
    125 mg/kg of body weight per day.

         The testicular toxicity of DCA was evaluated in adult male rats
    given both single and multiple (up to 14 days) oral doses. Delayed
    spermiation and altered resorption of residual bodies were observed in
    rats given single doses of 1500 or 3000 mg/kg of body weight; these
    effects persisted to varying degrees on post-treatment days 2, 14 and
    28. Delayed spermiation and formation of atypical residual bodies were
    also observed on days 2, 5, 9 and 14 in rats dosed daily with 54, 160,
    480 or 1440 mg/kg of body weight per day. Distorted sperm heads and
    acrosomes were observed in step 15 spermatids after administration of
    doses of 480 and 1440 mg/kg of body weight per day for 14 days.
    Decreases in the percentage of motile sperm occurred after 9 days at

    doses of 480 and 1440 mg/kg of body weight per day and after 14 days
    at 160 mg/kg of body weight per day. Increased numbers of fused
    epididymal sperm were observed on days 14, 9 and 5 in rats dosed with
    160, 480 or 1440 mg/kg of body weight per day, respectively; other
    morphological abnormalities occurred at 160 mg/kg of body weight per
    day and higher. On day 14, a significant decrease in epididymis weight
    was observed at 480 and 1440 mg/kg of body weight per day, and
    epididymal sperm count was decreased at 160 mg/kg of body weight per
    day and higher. These studies demonstrate that the testicular toxicity
    induced by DCA is similar to that produced by the analogue DBA (see
    section 4.2.3.2). However, the testicular toxicity of DCA is less
    severe at equal mol/litre concentrations. Moreover, the DCA-induced
    testicular lesions occur at lower doses as the duration of dosing
    increases, indicating the importance of using low-dose subchronic
    exposures to assess the health risk of prevalent DBPs (Linder et al.,
    1997a).

         DCA was found to be more potent than TCA in inhibiting
     in vitro fertilization of B5D2F1 mouse gametes (Cosby & Dukelow,
    1992). For DCA, the percentage of gametes fertilized dropped from
    87.0% to 67.3% at the lowest concentration tested, 100 mg/litre. No
    effects were noted for TCA at 100 mg/litre; at 1000 mg/litre, however,
    71.8% of the gametes were fertilized. Both responses were
    statistically different from those of their concurrent control groups.

    4.2.1.3  Developmental effects

         DCA has been shown to induce soft tissue abnormalities in fetal
    rats when administered by gavage in a water vehicle to their dams
    during gestation days 6-15 (Smith et al., 1992). These effects were
    observed at doses of 140 mg/kg of body weight per day and above and
    were not observed at 14 mg/kg of body weight per day. The heart was
    the most common target organ. An interventricular septal defect
    between the ascending aorta and the right ventricle was most commonly
    observed. Urogenital defects (bilateral hydronephrosis and renal
    papilla) and defects of the orbit were also observed. In a subsequent
    publication, these authors identified the most sensitive period to be
    days 12-15 of gestation (Epstein et al., 1992).

         Some limited experimentation has been conducted to evaluate the
    developmental effects of DCA in rat whole embryo culture (Saillenfait
    et al., 1995). The applicability of these data to the  in vivo 
    situation is difficult to evaluate based on the very high
    concentrations that were utilized in these studies. At a concentration
    of 1 mmol/litre, DCA retarded growth of the embryos by a variety of
    measures. It required 2.5 mmol/litre to induce brain and eye defects
    and 3.5 mmol/litre to produce abnormalities in other organ systems.

    4.2.1.4  Neurotoxicity

         Yount et al. (1982) demonstrated that DCA administered to rats in
    their feed at doses of 2.5-4 mmol/kg of body weight per day (322-716
    mg/kg of body weight per day) produced hindlimb weakness and abnormal
    gait within 2-4 weeks of treatment. These effects were associated with

    significant reductions in nerve conduction velocity and a decrease in
    the cross-sectional area of the tibial nerve.

    4.2.1.5  Toxicity in humans

         DCA was first investigated as a potential orally effective
    hypoglycaemic agent. Stacpoole et al. (1978) found that DCA,
    administered in doses of 3-4 g of dichloroacetate as the sodium salt
    per day for 6-7 days (43-57 mg/kg of body weight per day for a 70-kg
    person), significantly reduced fasting hyperglycaemia in patients with
    diabetes mellitus alone or in combination with hyperlipoproteinaemia
    by 24%. In addition, plasma lactic acid concentrations dropped by 73%
    and plasma alanine concentrations by 82%. Only mild sedation was
    noticed by some of the patients, and there was no evidence of altered
    blood counts or prothrombin time. There was a slower, but significant,
    decline in plasma triglyceride levels and less consistent effects on
    plasma cholesterol. -Hydroxybutyrate concentrations in plasma
    increased significantly and continuously over the 6 days of
    administration. Uric acid concentrations in serum increased and those
    in urine decreased, reflecting a 50% decrease in urinary clearance.

         Further details on the early investigations of DCA as an oral
    antidiabetic agent have been reviewed extensively (Crabb et al., 1981;
    Stacpoole, 1989; Stacpoole & Greene, 1992) and will not be dwelt on
    extensively in the present review. The main reason that DCA was not
    fully developed for this application was that longer-term
    administration to patients induced a reversible polyneuropathy (Moore
    et al., 1979b; Stacpoole, 1989). As discussed above, similar pathology
    was reported in experimental animals by several authors, which added
    significant weight to this observation. Subsequent work in rats
    suggested that thiamine deficiency contributed to the development of
    peripheral neuropathy (Stacpoole et al., 1990). However, a recent
    report (Kurlemann et al., 1995) indicated that supplementing the diet
    with 100 mg of thiamine daily did little to ameliorate the development
    of polyneuropathy of a patient treated with DCA at a dose of 100 mg/kg
    of body weight for 20 weeks.

         Subsequently, DCA has been investigated extensively in the
    treatment of congenital lactic acidosis (Coude et al., 1978; Stacpoole
    & Greene, 1992; Toth et al., 1993), a disease that is frequently
    fatal. More recently, DCA has been evaluated with success in the
    treatment of lactic acidosis associated with severe malaria in
    children (Krishna et al., 1995). It is also apparently effective in
    treatment of lactic acidosis associated with liver transplantation
    (Shangraw et al., 1994). DCA has been reported to improve psychiatric
    symptoms, as well as to markedly decrease elevated lactic acid levels
    in an individual with mitochondrial myopathy (Saijo et al., 1991).
    Doses of up to 100 mg/kg of body weight per day were administered. In
    a second study, DCA was shown to reverse lesions to the basal ganglia
    in a patient with complex I deficiency as measured by computerized
    tomography (CT) scan and magnetic resonance imaging (MRI) (Kimura et
    al., 1995). However, a second patient with pyruvate dehydrogenase
    complex deficiency displayed only transient improvement.

         Based on work done in animals, several studies have examined the
    effects of DCA on cardiovascular function under a variety of disease
    conditions and altered physiological states. DCA was found to
    stimulate myocardial lactate consumption and improve left ventricular
    efficiency in 10 patients with congestive heart failure (Bersin et
    al., 1994). On the other hand, although DCA decreased blood lactate
    levels in patients with congestive heart failure, it had no effect on
    exercise time, peak exercise oxygen consumption or flow to the
    exercising leg (Wilson et al., 1988). In normal individuals, DCA
    decreased blood lactate levels when exercising at less than 80% of
    average maximal oxygen consumption, but it did not affect blood
    lactate concentrations at exhaustion (Carraro et al., 1989).

         These studies show that DCA has some beneficial effects in a
    variety of metabolic diseases. Acutely, DCA produces little in the way
    of risk, because short exposures are without apparent adverse effect.
    However, despite the limited number of subjects that have been
    studied, the results, when coupled with the results of animal
    experiments, indicate relatively strongly that DCA is neurotoxic in
    humans. The delayed induction of these toxicities may be attributable
    in part to the fact that systemic concentrations of DCA can be
    expected to sharply increase with prolonged treatment, at least at the
    high doses used therapeutically, typically in the range 25-100 mg/kg
    of body weight. There is no human evidence of adverse effects at lower
    exposures to DCA (e.g., those that would be derived from drinking
    chlorinated water).

    4.2.1.6  Carcinogenicity and mutagenicity

         IARC has evaluated the carcinogenicity of DCA and, based on the
    data available at the time, concluded that there is inadequate
    evidence for its carcinogenicity in humans and limited evidence for
    its carcinogenicity in experimental animals. The compound was assigned
    to Group 3: not classifiable as to its carcinogenicity to humans
    (IARC, 1995).

         DCA is a very effective inducer of hepatic tumours in both mice
    and rats at high doses. Several studies in male and female B6C3F1
    mice found multiple tumours per animal with treatment concentrations
    of 2 g/litre and above with as little as 1 year of treatment
    (Herren-Freund et al., 1987; Bull et al., 1990; DeAngelo et al., 1991;
    Daniel et al., 1992a; Pereira, 1996). These studies are summarized in
    Table 15. Early in treatment (i.e., 52 weeks), the dose-response curve
    is very steep, with essentially no response observed at concentrations
    of 1 g/litre, but as many as four tumours per liver in mice treated
    with 2 g/litre (Bull et al., 1990). However, concentrations as low as
    0.5 g/litre will result in a hepatic tumour incidence of approximately
    80% in a full 2-year study (Daniel et al., 1992a). 

         Hepatic tumours are also induced by DCA in male F344 rats
    (Richmond et al., 1995). High doses of DCA given to rats also produce
    overt signs of peripheral neuropathy. Nevertheless, increased
    incidences of hyperplastic nodules, hepatocellular adenoma and
    hepatocellular carcinoma were observed at 60 weeks of treatment at


        Table 15. Carcinogenic effects of dichloroacetate in rodents

                                                                                                                                      
    Species        Dose        Duration   Tumour             HN & HAa                      HCb                 Reference
    (sex)          (g/litre)   (weeks)    site                                                             
                                                    Incidence   Tumour/n         Incidence    Tumour/n
                                                                (multiplicity)                (multiplicity)
                                                                                                                                      

    Mice

    B6C3F1 (M)     0           61                                                                              Herren-Freund et al. 
                   5           61         Liver     25/26       (4.6)            21/26        (1.7)            (1987)

    B6C3F1 (M)     1           52         Liver     2/11        0.3              -            -                Bull et al. (1990)
                   2           52         Liver     23/24       3.6              5/24         0.25
                   2           37         Liver     7/11        2.2              0/11         0

    B6C3F1 (M)     0           60         Liver     0/10        0                0/10         0                DeAngelo et al. (1991)
                   0.5         60         Liver
                   3.5         60         Liver     12/12       2.3              8/12         1.7
                   5           60         Liver     27/30       2.3              25/30        2.2
                   0           75         Liver     2/28        0.07
                   0.05        75         Liver     4/29        0.31
                   0.5         75         Liver     3/27        0.11
                   0           104        Liver     1/20        0.05             2/20         0.1              Daniel et al. (1992a)
                   0.5         104        Liver     12/24       0.5              15/24        0.63

    B6C3F1 (F)     0           52         Liver     1/40        0.03             0/40         0                Pereira (1996)
                   0.28        52         Liver     0/40        0                0/40         0
                   0.93        52         Liver     3/20        0.20             0/20         0
                   2.8         52         Liver     7/20        0.45             1/20         0.1
                   0           81         Liver     2/90        0.02             2/90         0.02
                   0.28        81         Liver     3/50        0.06             0/50         0
                   0.93        81         Liver     7/28        0.32             1/28         0.04
                   2.8         81         Liver     16/19       5.6              5/19         0.37
                                                                                                                                      

    Table 15. (continued)

                                                                                                                                      
    Species        Dose        Duration   Tumour            HN & HAa                       HCb                 Reference
    (sex)          (g/litre)   (weeks)    site                                                             
                                                    Incidence   Tumour/n         Incidence    Tumour/n
                                                                (multiplicity)                (multiplicity)
                                                                                                                                      

    Rats

    F344 (M)       0           60         Liver     0/7         0                0/7          0                Richmond et al. (1995)
                   0.05        60         Liver     0/7         0                0/7          0
                   0.5         60         Liver     0/7         0                0/7          0
                   2.4         60         Liver     (26/27)     0.96             1/27         0.04
                   0           104        Liver     1/23        0.04             0/23         0
                   0.05        104        Liver     0/26        0                0/26         0
                   0.5         104        Liver     (9/29)      0.31             3/29         0.1
                   2.4         104        Liver     NRc         NR               NR           NR

    F344 (M)       0           104        Liver     1/33        0.03             1/33         0.03             DeAngelo et al. (1996)
                   0.05        104        Liver     0/26        0                0/26         0
                   0.5         104        Liver     5/29        0.17             3/29         0.10
                   1.6d        104        Liver     4/28        0.14             6/28         0.24
                                                                                                                                      

    a  Combined hepatocellular nodules and hepatocellular adenomas.
    b  Hepatocellular carcinoma.
    c  NR = not reported.
    d  Concentration was 2.6 g/litre of drinking-water for 18 weeks and then lowered to 1 g/litre to give a mean daily concentration 
       of 1.6 g/litre.
    

    2.4 g/litre (Table 15). As in mice, if DCA treatment was extended to
    104 weeks, the incidence of these lesions was 41% in a group of 29
    rats at a treatment concentration of 0.5 g/litre. No tumours were
    observed at 0.05 g/litre, and only one hepatic tumour was observed in
    23 control rats.

         Estimated doses of DCA in mg/kg of body weight per day for the
    studies in Table 15 were as follows (ILSI, 1997; US EPA, 1998a):

         Herren-Freund et al. (1997): 5 g/litre = 1000 mg/kg of body
              weight per day
         Bull et al. (1990): 1 or 2 g/litre = 140 or 300 mg/kg of body
              weight per day
         DeAngelo et al. (1991): 0.05, 0.5, 3.5 or 5 g/litre = 7.6, 77,
              410 or 486 mg/kg of body weight per day
         Daniel et al. (1992a): 0.5 g/litre = 95 mg/kg of body weight per
              day
         Pereira (1996): 0.26, 0.86 or 2.6 g/litre = 40, 120 or 330 mg/kg
              of body weight per day
         Richmond et al. (1995): 0.05, 0.5 or 2.4 g/litre = 4, 40 or
              300 mg/kg of body weight per day
         DeAngelo et al. (1996): 0.05, 0.5 or 1.6 g/litre = 4, 40 or
              140 mg/kg of body weight per day.

         Male F344 rats were exposed for 2 years to DCA in their
    drinking-water at concentrations of 0.05, 0.5 or 1.6 g/litre. Based
    upon the pathological examination, DCA induced observable signs of
    toxicity in the nervous system, liver and myocardium. However,
    treatment-related neoplastic lesions were observed only in the liver.
    A statistically significant increase in carcinogenicity
    (hepatocellular carcinoma) was noted at 1.6 g/litre. Exposure to 0.5
    g/litre increased hepatocellular neoplasia (carcinoma and adenoma) at
    100 weeks. Calculation of the time-weighted mean daily dose at which
    50% of the animals exhibited liver neoplasia indicated that the F344
    male rat (approximately 10 mg/kg of body weight per day) is 10 times
    more sensitive than the B6C3F1 male mouse (approximately 100 mg/kg of
    body weight per day) (DeAngelo et al., 1996). 

         The ability of DCA to induce damage to DNA that could give rise
    to mutations or chromosomal damage has been studied both  in vivo and
     in vitro. Classical evaluations of DCA in  Salmonella typhimurium 
    tester strains, both with and without metabolic activation, have been
    largely negative if held to the standard of at least a 2-fold increase
    in apparent mutation frequency (Waskell, 1978; Herbert et al., 1980).
    However, a number of more recent studies have suggested some potential
    for DCA-induced modifications in DNA. DeMarini et al. (1994) reported
    that DCA induced prophage in  Escherichia coli at a concentration of
    0.26 mmol/litre and produced 2.7 and 4.2 revertants per ppm in
     S. typhimurium strain TA100 with and without S9 addition,
    respectively. There are some difficulties in interpreting this report,
    as the authors introduced DCA as a vapour, and it is not clear whether
    the concentrations reported (i.e., ppm) refer to air or medium
    concentrations. Second, at least in the case of the  Salmonella 
    assay, the DCA was introduced as the free acid and allowed to vaporize

    and partition into the incubation medium. Because DCA is a strong
    acid, and if sufficient time is allowed, such conditions could result
    in near-quantitative transfer of DCA to the medium. The amount
    volatilized in this case was approximately 60-600 mmol. Therefore, it
    is likely that the pH of this small amount of medium (2.5 ml) was
    substantially modified, even if only a fraction of this relatively
    large amount of strong acid was indeed transferred to the medium. The
    amount of DCA introduced into the prophage assay was unclear because
    the method of addition was not described, although the introduction to
    the journal article implied that it was again being tested as a
    volatile.

         Fox et al. (1996) recently published an evaluation of the
    mutagenic effects of sodium dichloroacetate. These investigations
    found no evidence of increased mutation rates in  Salmonella 
     typhimurium tester strains TA98, TA100, TA1535 or TA1537;
     Escherichia coli strain WP2urvA; or the mouse lymphoma forward
    mutation assay, whether incubated in the presence or absence of rat
    liver S9 fraction for metabolic activation. These authors found no
    evidence that DCA was capable of inducing chromosomal aberrations in
    CHO cells  in vitro at doses of up to 1100 mg/kg of body weight for
    3 days. These studies utilized neutralized DCA, supporting the
    contention that positive results in prior studies may have been due to
    artefactual results obtained by testing of the free acid or because
    various sources of DCA have greater amounts of impurities.

         Giller et al. (1997) examined the mutagenicity of DCA in the SOS
    chromotest, the Ames fluctuation assay and the newt micronucleus
    assay. DCA induced a positive response at 500 g/ml (approximately 3.5
    mmol/litre) in the SOS chromotest and at concentrations ranging from
    100 to 1500 g/ml (approximately 1-10 mmol/litre) in the Ames
    fluctuation assay. The effects were observed at a lower concentration
    in the absence of S9. The concentrations used in these studies exceed
    the peak systemic concentrations of DCA that produce a high incidence
    of liver tumours in mice by approximately 3 orders of magnitude
    (Kato-Weinstein et al., 1998). Moreover, it appears that the authors
    utilized the free acid in these experiments, raising the possibility
    of a pH artefact. The newt micronucleus assay was found to be
    negative.

         DCA has been shown to produce a mutagenic and clastogenic
    response in the  in vitro mouse lymphoma assay, but only at doses at
    or above 1 mmol/litre (Harrington-Brock et al., 1998).

         Analogous difficulties have been encountered when attempting to
    document the mutagenic effects of DCA  in vivo. Nelson & Bull (1988)
    and Nelson et al. (1989) reported that DCA induced single strand
    breaks (SSB) in hepatic DNA when administered by gavage to both mice
    and rats. Subsequent investigators were unable to replicate these
    results in detail (Chang et al., 1992; Daniel et al., 1993a). However,
    a small transitory increase in SSB was observed with doses of 5 and
    10 mmol/kg of body weight in male B6C3F1 mice (Chang et al., 1992).
    The bases of the discrepancies in these results are not clear, but
    could, in part, be attributed to slightly different methods. As noted

    in the subsequent section on TCA (section 4.2.2), the Nelson & Bull
    (1988) results were not replicated by Styles et al. (1991), although
    there was greater similarity in the methods used. More recently,
    Austin et al. (1996) showed that acute doses of DCA oxidatively damage
    nuclear DNA, measured as increases in the 8-hydroxy-2-deoxyguanosine
    (8-OH-dG) relative to 2-deoxyguanosine content of the isolated DNA.
    The time course of this damage is more consistent with the development
    of SSB breaks reported by Chang et al. (1992) and could represent the
    repair process that involves strand scission. There are two important
    points that must be made: (i) the induction of SSB by Chang et al.
    (1992) was very small relative to that seen with the positive
    controls, diethylnitrosamine and methylmethane sulfonate; and (ii)
    although increased 8-OH-dG was observed with acute treatments with
    DCA, there was not a sustained elevation of this adduct in nuclear DNA
    of mice when treatments were extended to 3 or 10 weeks in
    drinking-water (Parrish et al., 1996).

         Fuscoe et al. (1996) reported results obtained with the mouse
    peripheral blood micronucleus assay. They found a small, but
    statistically significant, increase in polychromatic erythrocytes
    containing micronuclei in male B6C3F1 mice treated for 9 days with
    3.5 g of DCA per litre of drinking-water. However, this response was
    not maintained through 28 days of exposure. These investigators also
    examined DNA migration in the single-cell gel assay. In this case, DCA
    appeared to retard migration of DNA, suggesting the possibility of DNA
    cross-linking after 28 days of treatment at 3.5 g/litre. Neither assay
    revealed significant effects of DCA at concentrations of 2 g/litre or
    below. DCA induces 3-4 tumours per animal within 1 year at 2 g/litre
    in drinking-water (Bull et al., 1990). The higher dose adds little to
    the tumorigenic response. More information with regard to the
    possibility that DCA can cause mutations in liver cells is found in a
    recent study using the  lacI locus in the Big Blue(R) transgenic
    mouse mutagenesis assay (Leavitt et al., 1997). These investigators
    used a drinking-water route and the same doses of DCA as were used in
    the rodent bioassay. After 10 and 60 weeks of DCA administration, an
    increased frequency of mutants was observed at the high dose (3.5
    g/litre). Mutational spectral analysis of these mutations revealed a
    different spectrum in the mutants from DCA-treated animals than was
    seen in the untreated animals. At this high dose of DCA, a large
    portion of the liver can actually be tumour tissue. Because tumours
    result from clonal expansion, the presence of tumour tissue in the
    evaluated sample would give a falsely high mutation frequency if a
     lacI mutation occurred in the rapidly expanding tumour clone. Thus,
    these indications of genotoxic activity may have little to do with the
    induction of hepatic cancer by DCA.

         DCA appears to specifically stimulate outgrowth of hepatocellular
    adenomas, rather than hepatocellular carcinomas. Pereira & Phelps
    (1996) examined the role of DCA as a promoter of methylnitrosourea
    (MNU)-initiated hepatic tumours in female B6C3F1 mice. These data are
    provided in graphic form in Figure 2. At a concentration of
    2.6 g/litre of drinking-water, DCA induced a very large increase in
    the number of hepatocellular adenomas, but had no significant effect
    on the induction of hepatocellular carcinomas. These data would appear

    to be consistent with the stop experiments of Bull et al. (1990), who
    found that suspension of treatment with DCA appeared to arrest
    progression of liver tumours, but resulted in a yield of
    hepatocellular adenomas and nodules that was proportional to the total
    dose of DCA administered. In contrast, most of the tumours that
    remained after the suspension of TCA treatment for 3 months were
    hepatocellular carcinomas.

         More recent studies on the effects of DCA on cell replication
    within normal hepatocytes and hyperplastic nodules and tumours
    (predominantly adenomas) indicate that DCA has selective effects.
    Stauber & Bull (1997) found that DCA had a small, stimulatory effect
    on the replication rate of normal hepatocytes over the first 14 days
    of treatment. As treatment was extended to 28 days and beyond, these
    effects became inhibitory at concentrations in drinking-water of
    0.5 g/litre and above. In contrast, hepatocytes within nodules and
    tumours appeared to be resistant to the inhibitory effects of DCA. At
    a concentration of 2 g/litre, DCA doubled the rate at which c-Jun
    immunoreactive hepatocytes replicated within hyperplastic nodules and
    adenomas. This strong stimulation of tumour cell replication would
    appear to be responsible for the very rapid induction of tumours in
    mice treated with DCA in drinking-water at concentrations of 2 g/litre
    and above. It would appear that the slower induction of liver tumours
    at lower doses of DCA depends primarily on the selective suppression
    of the replication of normal hepatocytes relative to that of initiated
    cells.

         The inhibitory effect of DCA on replication of normal hepatocytes
    has been observed by a number of investigators (Carter et al., 1995).
    The rate of replication is sharply inhibited within 5 days at
    concentrations of DCA of 5 g/litre. At 0.5 g/litre, the replication
    rate becomes inhibited to the same extent as observed with 5 g/litre
    after 20 days of treatment. These decreases in replication were
    accompanied by an increase in the percentage of the cells that were
    mononucleated, which is probably associated with an increase in
    tetraploid cells.

         The suppression of cell replication by DCA in normal hepatocytes
    of treated mice is accompanied by decreases in apoptosis (Snyder et
    al., 1995). At concentrations of 5 g/litre, the frequency at which
    apoptotic cells are observed drops by 60-75% with as few as 5 days of
    treatment. At 0.5 g/litre, there is a downward trend that is observed
    over the period from 5 to 30 days such that the frequency of apoptotic
    bodies at this low dose approaches that observed at the highest dose
    at 30 days. This result essentially parallels that described above for
    suppression of the rates of cell replication. This raises a dilemma as
    to whether the driver of the response is suppressed replication or
    suppressed apoptosis. Whichever is the case, this has to translate
    into suppressed turnover of normal hepatocytes. The question is
    whether this suppressive effect on cell turnover increases the
    probability of transformation of hepatocytes.

    FIGURE 2

         Small, but statistically significant, increases in the rate of
    DNA synthesis in primary cultures of rat hepatocytes have been
    reported at a concentration of 1 mmol of DCA per litre (Reddy et al.,
    1992). This is indirect evidence of an effect on the rate of cell
    division, because replication rates were not actually measured in this
    study. Nevertheless, it seems probable that DCA does act as a weak
    mitogen. Adaptation or down-regulation of this response has been
    consistently observed  in vivo as described above. The key
    observation appears to be that the mitogenic response is not
    down-regulated in hyperplastic nodules or tumours (Stauber & Bull,
    1997).

         Hyperplastic nodules and tumours induced by DCA have some common
    characteristics that distinguish them from nodules and tumours that
    are induced by TCA. In female mice, Pereira (1996) indicated that
    liver tumours induced by DCA tended to be eosinophilic, whereas those
    induced by TCA were basophilic. In male B6C3F1 mice treated with 2 g
    of DCA per litre, a substantial fraction (66%) of the altered hepatic
    foci found and nodules were reported to be eosinophilic. However, the
    larger lesions tend to be basophilic (Stauber & Bull, 1997). These
    larger lesions included hyperplastic nodules, adenomas and carcinomas.
    These data suggest that there are some differences in tumour induction
    by DCA based on sex. However, this difference appears to be important
    primarily at high doses (>2 g/litre), where the rate of cell
    replication is enhanced in a set of basophilic lesions. The
    development of these lesions may account for the much shorter
    latencies observed in male mice as compared with female mice at high
    doses.

         As pointed out by previous investigators examining responses in
    male mice (Bull et al., 1990; DeAngelo et al., 1991), Pereira (1996)
    found the dose-response curves describing the induction of total
    lesions by DCA to be non-linear in female mice. Conversely, the
    effects of TCA are essentially linear with dose.

         Stauber & Bull (1997) found that DCA-induced liver tumours in
    male mice were immunoreactive to c-Jun and c-Fos antibodies, whereas
    TCA-induced liver tumours were not. This difference would appear
    consistent with the observation that DCA-induced tumours in female
    mice expressed the GST-pi at high levels, whereas TCA-induced tumours
    were largely GST-pi negative. The expression of GST-pi is dependent on
    AP-1 transcription factor binding sites in the promoter region of the
    gene. Thus, elevations of c-Jun and c-Fos would be expected to
    increase GST-pi expression (Angel & Karin, 1991). Conversely,
    peroxisome proliferator activated receptor (PPAR)-alpha is known to
    interfere with the c-Jun activity (Sakai et al., 1995). As a
    consequence, GST-pi is generally not observed in tumours induced by
    peroxisome proliferators.

         Tao et al. (1996) reported a further differentiation of DCA- and
    TCA-induced tumours. Non-neoplastic hepatocytes observed in mice
    treated with DCA were found to have high levels of TGF-alpha, whereas
    cells within the tumour expressed much lower levels. The opposite was
    observed with TGF- expression, which was high in tumours and low in

    normal tissues. This differential distribution of expression was not
    observed in non-involved tissue and tumours from TCA-treated mice. The
    precise involvement of these growth factors in the growth and
    development of tumours cannot be stated. However, both TGF-alpha and
    TGF- are known to be intimately involved with cell birth and cell
    death processes. In liver tissue, TGF- expression is associated with
    apoptosis (programmed cell death) and TGF-alpha expression is
    associated with proliferative states. It should be noted that it is
    not known whether these differences reflect characteristics of the
    neoplastic cells or are actually responses induced by DCA.

         Anna et al. (1994) and Ferreira-Gonzalez et al. (1995)
    independently assessed the frequency and spectra of H- ras mutations
    in DCA-induced tumours. These data and those of historical controls
    for male B6C3F1 mice (Maronpot et al., 1995) specifically at codon-61
    of H- ras are displayed in Table 16. The mutation frequency in
    DCA-induced tumours does not differ significantly from that observed
    in spontaneous tumours. However, there is an obvious change in the
    mutation spectra in codon 61, involving a significant increase in the
    H- ras-61(CTA) mutation largely at the expense of the H- ras-61(AAA)
    lesion. A traditional interpretation of changes in mutation spectra
    would be that this is evidence of mutation (Reynolds et al., 1987). As
    pointed out by Anna et al. (1994), however, such an effect could be
    accounted for if cells expressing a particular mutation were selected
    for by treatment. Since the H- ras-61(CTA) mutation codes for
    leucine, a neutral amino acid, whereas the H- ras-61(AAA) mutation
    codes for lysine, a charged amino acid in this position, the
    structures of these two mutant proteins are potentially quite
    different in the Switch 2 region of the  ras protein. Alterations in
    structure within this region could significantly affect the affinity
    of H- ras binding to  raf-1 and other proteins involved in signal
    transduction (Drugan et al., 1996).

    4.2.1.7  Comparative pharmacokinetics and metabolism

         The mammalian metabolism of DCA has received relatively little
    study. However, there is sufficient information to show that its
    metabolism is very dose-dependent and is dramatically affected by
    prior exposure. While some significant differences in the details of
    metabolism appear between species, these general statements hold for
    both rodents and humans.

         A proposed metabolic scheme for DCA, adapted from Larson & Bull
    (1992), is provided in Figure 3. Oxalate, glyoxylate, MCA and carbon
    dioxide have all been established as metabolites of DCA (Stacpoole,
    1989; Larson & Bull, 1992; Lin et al., 1993; Gonzalez-Leon et al.,
    1997). In addition to the metabolites depicted, thiodiacetate has been
    observed in small amounts in the urine of mice and rats (Larson &
    Bull, 1992). This may arise from the reaction of MCA with GSH (Yllner,
    1971), but other mechanisms are also possible. The intermediates
    indicated are hypothetical but reasonable in terms of the end-products
    observed. The extent to which reductive dehalogenation and peroxy
    radical formation play a role in the metabolism of DCA is unclear. As
    discussed later, such reactions clearly play a role in the metabolism


        Table 16. Mutation frequency and spectra with codon 61 of H- ras of B6C3F1 mice treated with dichloroacetate 
              and trichloroacetatea

                                                                                                                      
    Chemical                      No. of H-ras 61/   Fraction     CAA          AAA           CGA          CTA
                                  no. of tumours
                                                                                                                      

    Spontaneous hepatocellular 
      carcinomasb                 183/333            0.56         150 (0.45)   106 (0.32)    50 (0.15)    21 (0.06)
    Dichloroacetateb,c,d          61/110             0.55         48 (0.44)    15 (0.14)     25 (0.22)    22 (0.20)
    Trichloroacetatec             5/11               0.45         6 (0.55)     4 (0.36)      1 (0.09)     0 (0)
                                                                                                                      

    a  Mutations at other codons are not included, although these tumours are kept as part of the denominator. 
       Therefore, all mutants and wild-type at codon 61 do not add up to the total number of tumours.
    b  Anna et al. (1994).
    c  Ferreira-Gonzalez et al. (1995).
    d  Maronpot et al. (1995).
    

    of trihaloacetates, and they are included here for completeness. An
    additional pathway to glycolate could also be rationalized by the
    oxidative metabolism of MCA. 

         The human metabolism of DCA first came under study because of its
    proposed use as an oral hypoglycaemic agent. Lukas et al. (1980)
    studied intravenously infused (over a 20-min interval) doses of 10 and
    20 mg/kg of body weight in two human volunteers at each dose. The low
    and high doses led to mean half-lives of 0.34 and 0.51 h,
    respectively. Wells et al. (1980) noted that the peak concentration
    increased disproportionately when intravenous doses (30-min infusion)
    of DCA increased from 1 to 50 mg/kg of body weight and departed from
    linearity as doses approached 30 mg/kg of body weight. Whereas the
    half-life of DCA in this study was seen to be approximately 20 min at
    doses below 25 mg/kg of body weight, the half-life at higher doses was
    closer to 40 min, and the mean half-life was 31.8  10.9 (standard
    deviation [SD]) for all 11 subjects. The authors also noted that the
    effects of DCA on plasma lactate and alanine persisted several days
    after cessation of repeated oral treatment with DCA, but ended within
    12 h after administering a single intravenous dose. Curry et al.
    (1985) found that the mean half-life of DCA increased from 63.3 min to
    an average of 374 min following the fifth of a series of 50 mg/kg of
    body weight doses of DCA administered intravenously at 2-h intervals.
    These data suggest that high repeated doses of DCA appear to hinder
    the metabolic clearance of DCA.

         There are substantial species differences in the metabolism of
    DCA. Dogs, in particular, clear DCA from blood at a very low rate.
    Lukas et al. (1980) found that the half-life of DCA administered as a
    single 100 mg/kg of body weight dose was between 17.1 and 24.6 h. In
    contrast, the clearance of the same dose in rats occurred with a
    half-life of 2.1-4.4 h. The very much lower metabolic clearance of DCA
    in the dog is probably responsible for its much greater acute toxicity
    in this species (Katz et al., 1981). However, the half-life of DCA in
    humans is much closer to that in rats than to that in dogs (Curry et
    al., 1991).

         Larson & Bull (1992) studied the metabolism of DCA in mice and
    rats. These authors estimated a half-life of 1.5 h in mice and 0.9 h
    in rats following oral doses. The estimates of half-life in mice in
    this study were problematic, because it was clear that there is a
    tremendous first-pass effect on DCA's absorption from the
    gastrointestinal tract, which was particularly marked in mice. These
    authors also estimated a maximum concentration of DCA from oral doses
    of 20 and 100 mg/kg of body weight. In mice, the  Cmax was found to
    be 4 and 20 nmol/ml, respectively. In rats, the  Cmax was found to
    be 15 and 380 at the same doses. These authors also found that
    significant amounts of DCA were metabolized to carbon dioxide and the
    non-halogenated acids -- glycolate, glyoxylate and oxalate. The
    fraction of DCA that was metabolized to carbon dioxide was
    substantially underestimated in these studies, as was shown in the
    more recent study of Xu et al. (1995), in which approximately 45% of
    an oral dose of DCA was metabolized to carbon dioxide in mice within
    24 h. Lin et al. (1993) also provided more definitive analyses of the

    FIGURE 3

    production of glycolate, glyoxylate and oxalate as major urinary
    metabolites of DCA in the F344 rat. It is notable for later
    discussions that these authors found a smaller percentage of the dose
    of 1-14C-DCA ending up in glycolate than was observed with 2-14C-DCA.
    This was offset by a somewhat greater yield (not statistically
    significant) of carbon dioxide from 1-14C-DCA than from 2-14C-DCA.
    These data would suggest that there are alternative pathways from DCA
    to carbon dioxide in the rat besides the conversion to glyoxylate.

         The decreases in DCA clearance with repeated doses appear to be
    largely due to the inactivation of one enzyme involved in its
    metabolism. There appear to be subtle differences in the metabolism of
    mice and rats that lead to different manifestations of this
    inhibition. In experiments that involved pretreatment of F344 rats at
    a level of 0.2 or 2 g/litre in their drinking-water for a 14-day
    period, it was found that the conversion of an oral dose of DCA to
    carbon dioxide was substantially inhibited. Similar pretreatment of
    male B6C3F1 mice did not affect carbon dioxide production from DCA.
    However, in both cases, it is apparent that the metabolic clearance of
    DCA from blood was affected by the DCA pretreatment. In rats, the
    kinetics of DCA disappearance were studied with intravenous dosing,
    which allows more precise definition of the kinetics. This
    pretreatment led to an increase in the half-life of DCA in the blood
    of rats from 2.4  0.8 h to 10.8  2.0 h when animals were
    administered a dose of 100 mg/kg of body weight intravenously. Oral
    dosing was used in mice, and the major impact of pretreatment was to
    increase the  Cmax from 2.6  2.6 g/ml in naive mice to 129.9 g/ml
    in mice that had DCA in their drinking-water at 2 g/litre until the
    prior day (16 h before administration of the test dose). Therefore,
    the phenomenon that DCA treatment inhibited its own metabolism, which
    was originally observed in humans, could be replicated in both mice
    and rats.

         The mechanism by which this tremendous change in metabolism
    occurs has not been established. It has been shown by Lipscomb et al.
    (1995) that the bulk of the metabolism of DCA occurs in cytosolic
    fractions. Very little DCA is metabolized in microsomes. It is
    apparent that the metabolism of DCA in cytosolic preparations from
    rodent liver is dependent upon nicotinamide cofactor and GSH. However,
    the metabolism is not mediated by enzymes that can be recovered on a
    GSH-sepharose column. Glutathione transferase activities towards
    chlorodinitrobenzene were observed in the column. Subsequent work has
    shown that it is the activity in the cytosol, however, that is
    eliminated by DCA treatment (Gonzalez-Leon et al., 1997). Tong et al.
    (1998) showed that a substantial fraction of DCA's metabolism is
    mediated by a novel GST, GST-zeta. This enzyme appears to be subject
    to autoinhibition by DCA.

    4.2.1.8  Mode of action

         Some indication of mechanisms by which DCA produces its effects
    can be gleaned from studies cited in previous sections of this
    document. While DCA produces many different types of toxicological
    effects (i.e., neurotoxicity, reproductive and developmental

    toxicities and carcinogenicity), there may be some features that are
    common to the mechanisms that produce all these effects. The
    particulars of those mechanisms remain to be established. What follows
    is a brief outline of what is known and some speculation on how new
    research may be applied to developing this information to better
    assess the risks associated with the production of DCA as a by-product
    of the chlorination of drinking-water.

         Given the lack of evidence that induction of DNA damage by DCA is
    involved in liver cancer induction, are there plausible alternative
    mechanisms that may be invoked? The most common alternative mode of
    action would be evidence that carcinogenic doses of DCA induce
    cytotoxic damage in the target organ, which leads to reparative
    hyperplasia. Although there is some evidence of single-cell necrosis
    with chronic exposure (Stauber & Bull, 1997) and infarcts are
    occasionally observed (Sanchez & Bull, 1990) in the livers of B6C3F1
    mice, this mode of action plays a negligible role at the lowest doses
    that induce liver cancer. Small and variable initial increases in
    replication rates of hepatocytes in treated mice are reversed with
    continued treatment, with the dominant effect becoming inhibition of
    replication within a 4- to 8-week period (Carter et al., 1995; Stauber
    & Bull, 1997). These observations indicate that necrosis followed by
    reparative hyperplasia do not explain the rapid carcinogenic responses
    to DCA.

         Additional data add support to the hypothesis that DCA does in
    fact act largely by a "non-genotoxic" mode of action. Early data
    indicated that DCA was capable of inducing peroxisome proliferation
    (Nelson & Bull, 1988; DeAngelo et al., 1989). While these effects have
    been observed by others, they are clearly of short duration,
    disappearing within a few months (DeAngelo et al., 1989). Moreover, it
    has become apparent that DCA induces hepatic tumours at dose rates
    that are significantly below those required to induce peroxisome
    proliferation (cf. DeAngelo et al., 1989 and Daniel et al., 1992a).

         New research has shown that DCA acts primarily to increase the
    growth rate of pre-initiated cells in the liver. Stauber et al. (1998)
    found that DCA induces growth of colonies on soft agar when cell
    suspensions were obtained from the liver of neonatal mice. The most
    impressive aspect of these studies was that the colonies expressed the
    same phenotype (c-Jun+) as was found in DCA-induced tumours  in 
     vivo. Similar experiments produced a c-Jun phenotype when TCA was
    incorporated into the soft agar. A concentration of 0.5 mmol of DCA
    per litre was required in the medium to produce a significant increase
    in the number of colonies formed in a 10-day interval when the cells
    were derived from naive mice. However, if mice were pretreated with
    0.5 g of DCA per litre in their drinking-water prior to the isolation
    of the cells, 0.02 mmol of DCA per litre was as effective as 0.5
    mmol/litre. Moreover, the yield of colonies was increased by this
    pretreatment. The increased sensitivity to DCA appears closely related
    to the finding that pretreating animals with low levels of DCA
    (<0.2 g/litre of drinking-water) reduces its metabolism by more than
    90% (Gonzalez-Leon et al., 1997). The increased number of colonies
    produced by pretreatment appears to be due to clonal expansion of

    these cells. It appears that this activity accounts for the tumours
    induced by DCA at higher doses (>2 g/litre) where blood
    concentrations are found to be in the 100-500 mol/litre range.
    However, blood concentrations ranging from 1 to 7 mol/litre produced
    in mice treated with 0.5 g/litre (Kato-Weinstein et al., 1998) result
    in an 80% incidence of liver tumours (Daniel et al., 1992a),
    suggesting that a second mechanism may be involved at lower doses.

         At all carcinogenic doses studied, DCA increases the deposition
    of glycogen in the liver (Kato-Weinstein et al., 1998). This suggests
    that DCA is modifying cell signalling pathways. Low intraperitoneal
    doses of DCA produce increases in serum insulin concentrations in
    response to glucose challenge (Kato-Weinstein et al., 1998). In
    contrast, decreases in serum insulin concentrations have been observed
    in mice chronically administered DCA at either 0.5 or 2 g/litre (Smith
    et al., 1997). However, these measurements were made during the
    daylight hours when both serum glucose and blood DCA concentrations
    are low. Clearly, the involvement of insulin in DCA-induced liver
    tumorigenesis needs to be studied further.

         Recent studies have provided more substantive evidence that DCA
    at very high levels possesses some ability to induce genotoxic
    effects. Of particular note are the studies of Harrington-Brock et al.
    (1998), who found significant increases in mutant frequencies in mouse
    lymphoma cells  in vitro at concentrations in excess of 1 mmol/litre.
    Similar potency of DCA has been observed in the Ames fluctuation assay
    by Giller et al. (1997). Previous studies were largely negative or
    made use of the free acid form of DCA. Brusick (1986) had previously
    documented that low pH produces increased evidence of genotoxic damage
    in cultured mammalian cells.

         Leavitt et al. (1997) reported increased recovery of mutant cells
    from the  lacI transgenic mouse with varying periods of treatment
    with DCA in drinking-water. Significant increases were observed only
    when mice had been treated with 3.5 g/litre for 60 weeks, but not for
    shorter time intervals. No significant increases were noted at 1
    g/litre. Although the authors took care to ensure that nodules and
    tumours were excluded from the sampling, Stauber & Bull (1997)
    demonstrated that there are numerous lesions that are smaller than
    nodules in B6C3F1 mice maintained on 2 g of DCA per litre for only
    40 weeks. It was inevitable that some of these microscopic lesions
    were included within the tissue samples described. Given the marked
    stimulation of cell replication that occurs within lesions in mice, it
    is not possible to determine if the effect reported by Leavitt et al.
    (1997) is due to mutagenic effects of DCA or its demonstrated ability
    to selectively stimulate the growth of tumour phenotype.

         Based on the available evidence, it is probable that genotoxic
    effects of DCA play little, if any, role in the induction of liver
    cancer in rodents at low doses. This conclusion is based on clear
    evidence that DCA is capable of acting as a tumour promoter and
    produces effects on cell replication or apoptosis at all carcinogenic
    doses (Snyder et al., 1995; Pereira & Phelps, 1996; Stauber & Bull,
    1997). The concentrations of DCA required to produce genotoxic effects

     in vitro and the blood levels necessary to detect minimal genotoxic
    effects  in vivo are 3 orders of magnitude higher than those
    necessary for induction of an 80% tumour incidence. The recent
    evaluation by ILSI (1997) also concluded that the mechanism by which
    DCA increased liver tumours was non-genotoxic. However, new data
    indicate that the actual mechanism is by tumour promotion rather than
    by cytotoxicity and reparative hyperplasia.

         The metabolism of DCA is very dose-dependent, with metabolism and
    clearance of the chemical being inhibited sharply with high dose
    rates. Most data now suggest that it is the parent compound that is
    responsible for the effects related to carcinogenicity. Thus, simply
    on the basis of considerations of target organ dosimetry, the effects
    of DCA would be predicted to increase sharply at chronic dosing levels
    that approach or exceed 30 mg/kg of body weight per day rather than
    being simple linear functions of dose. Unfortunately, the available
    data do not allow the systemic dose versus external dose relationships
    to be determined with any precision on the basis of current
    information. In rats, the full inhibition was observed at
    concentrations of DCA in drinking-water as low as 0.2 g/litre
    (Gonzalez-Leon et al., 1997). The minimum treatment level for DCA's
    inhibition of its own metabolism in mice has not been established.
    However, blood concentrations of DCA increase from 2-4 mol/litre when
    the mice are treated with 0.5 g/litre in drinking-water to
    approximately 300 mol/litre when the treatment concentration is
    2 g/litre (Kato-Weinstein et al., 1998). These are peak concentrations
    measured during the night when mice are consuming the DCA-containing
    water. This 100-fold increase in blood concentrations with a mere
    4-fold increase in dose undoubtedly contributes to the highly
    non-linear tumorigenic response for DCA reported previously (Bull et
    al., 1990). The effect in humans has been documented to occur with
    doses in a similar range, 30 mg/kg of body weight (Wells et al.,
    1980). The animal data need to be extended to lower dosing rates or
    the human treatments need to be extended in time to more precisely
    define the level of chronic exposure that is required to produce this
    phenomenon. Alternatively, a better understanding of the mechanism of
    this inhibition should allow the critical question of whether the
    inhibition is a function of cumulative dose or daily dose to be
    answered.

         The available data indicate that DCA differentially affects the
    replication rates of normal hepatocytes and hepatocytes that have been
    initiated. The dose-response relationships are complex, with DCA
    initially stimulating division of normal hepatocytes. However, at the
    lower chronic doses used in animal studies (but still very high
    relative to those that would be derived from drinking-water), the
    replication rate of normal hepatocytes is eventually sharply
    inhibited. This indicates that normal hepatocytes eventually
    down-regulate those pathways that are sensitive to stimulation by DCA.
    However, altered cells, particularly those that express high amounts
    of a protein that is immunoreactive to a c-Jun antibody, do not seem
    to be able to down-regulate this response. Thus, the rates of
    replication in the preneoplastic lesions with this phenotype are very
    high at the doses that cause DCA tumours to develop with a very low

    latency. Preliminary data suggest that this continued alteration in
    cell birth and death rates is also necessary for the tumours to
    progress to malignancy (Bull et al., 1990). This interpretation is
    supported by studies that employ initiation/promotion designs as well
    (Pereira, 1996).

         On the basis of the above considerations, it is suggested that
    the currently available cancer risk estimates for DCA should be
    modified by incorporating newly developing information on its
    comparative metabolism and modes of action to formulate a biologically
    based dose-response model. These data are not available at the time of
    this writing, but should become available within the next 2-3 years.

         The dose-response data for effects other than cancer vary
    significantly, with dogs being extraordinarily sensitive (Katz et al.,
    1981). However, inhibition of the metabolism of DCA in chronically
    treated rodents (Gonzalez-Leon et al., 1997) and humans (Wells et al.,
    1980; Curry et al., 1985) may cause differences in sensitivities
    between species to converge somewhat with repeated treatment or
    exposure. However, a significant difference in species sensitivity
    remains. Cicmanec et al. (1991) identified a LOAEL of 12.5 mg/kg of
    body weight per day in dogs treated for 90 days. NOAELs for
    hepatomegaly in mice appear to be in the neighbourhood of 0.2 g/litre,
    which is approximately 40 mg/kg of body weight per day. On the basis
    of the rates of metabolic clearance and the assumption that the
    intrinsic sensitivities of different species are similar, the average
    human would seem to more closely approximate rats than dogs. To obtain
    more accurate pictures of human sensitivity at low doses, however, it
    is clear that future work must focus more specifically on
    toxicodynamic variables. 

    4.2.2  Trichloroacetic acid (trichloroacetate)

         Like DCA, TCA exists almost exclusively in the salt form at pHs
    found in drinking-water because of its very low p Ka of 0.70 (IARC,
    1995).

    4.2.2.1  General toxicological properties and information on 
             dose-response in animals

    1)   Acute toxicity

         Very little information exists on the mammalian toxicology of TCA
    before it was discovered as a by-product of drinking-water
    chlorination. Woodard et al. (1941) determined the oral LD50 of
    trichloroacetate (i.e., neutralized to pH 6-7) to be 3.32 g/kg of body
    weight in rats and 4.97 g/kg of body weight in albino mice when the
    compound was administered in aqueous solution. These values were in
    the same general range as those for neutralized acetic acid.

         Davis (1990) examined the effects of TCA on blood glucose and
    lactate levels following a dosing regimen of total doses of 0.92 or
    2.45 mmol/kg of body weight administered 3 times in 1 day to
    Sprague-Dawley rats by gavage. A typographical error in dosage was

    suspected and confirmed with the author (M.E. Davis, personal
    communication, 1996). Reductions in plasma glucose concentrations were
    observed in females at the high dose, and lactic acid levels were
    decreased at 0.92 and 2.45 mmol/kg of body weight doses in females,
    but only at the high dose in males. The authors noted that these high
    concentrations were neutralized with sodium hydroxide. However, these
    are very low doses of TCA relative to those found to produce similar
    effects in other studies. These initial experiments were followed by a
    study of the effects of TCA administered in drinking-water for 14 days
    at 0.04, 0.16, 0.63 or 2.38 g/litre. Effects on urine volume and
    osmolality were reported to occur at the highest dose, but not at
    0.63 g/litre. Effects on glucose and lactate were not reported. 

    2)   Short-term toxicity

         Mather et al. (1990) administered TCA to male Sprague-Dawley rats
    in drinking-water at concentrations of 0, 50, 500 or 5000 mg/litre (0,
    4.1, 36.5 or 355 mg/kg of body weight per day) for 90 days. Small, but
    statistically insignificant, decreases in body weight were observed at
    the highest dose. TCA produced a significant increase in the liver to
    body weight ratio at this dose, but not at 500 mg/litre. This effect
    was associated with a small, but statistically significant, increase
    in cyanide-insensitive acyl CoA oxidase activity in the liver, an
    indicator of peroxisome proliferation.

         Unlike DCA and MCA, TCA does not appear to be a substrate for the
    mitochondrial pyruvate carrier (Halestrap, 1975). TCA does appear to
    inhibit the pig heart pyruvate dehydrogenase kinase at approximately
    the same concentrations as for DCA (Whitehouse et al., 1974). However,
    the authors noted that DCA influenced the proportion of active
    pyruvate dehydrogenase in the perfused rat heart, but TCA was inactive
    under these circumstances. Although these data were obtained from
    mitochondria from different sources, they suggest that there may be
    some differences in effects of DCA and TCA on intermediary metabolism
    related to their transport into various cellular compartments.

         Bhat et al. (1991) administered TCA to male Sprague-Dawley rats
    at a concentration of 45.8 mmol/litre (7.5 g/litre) in their
    drinking-water for 90 days to provide an approximate intake of 785
    mg/kg of body weight per day. These levels produced minimal evidence
    of liver toxicity by histopathological examination. However, lower
    concentrations in drinking-water have been shown to seriously impair
    water and food consumption in experimental animals (Bull et al., 1990;
    Davis, 1990; Mather et al., 1990; DeAngelo et al., 1991).
    Consequently, it is difficult to determine how these data relate to
    the potential effects of the low concentrations of TCA that are found
    in chlorinated drinking-water (e.g., in the 10-100 g/litre range).

         The most obvious target organ for TCA is the liver. This effect
    is marked by a hepatomegaly (Goldsworthy & Popp, 1987; Bull et al.,
    1990; Mather et al., 1990; Sanchez & Bull, 1990), which is presumably
    related to the ability of TCA to act as a peroxisome proliferator
    (DeAngelo et al., 1989), since that is a common finding with this
    class of rodent carcinogens. It could also be related to the

    metabolism of TCA to DCA (Larson & Bull, 1992). However, more recent
    data suggest that the apparent conversion of TCA to DCA has largely
    been the result of artefactual conversion when fresh oxygenated blood
    is acidified prior to derivatization (Merdink et al., 1998). TCA is
    clearly without substantive cytotoxic effects at doses of less than
    300 mg/kg of body weight  in vivo (Bull et al., 1990; Sanchez & Bull,
    1990; Acharya et al., 1995) or concentrations of up to 5 mmol/litre
     in vitro (Bruschi & Bull, 1993).

    4.2.2.2  Reproductive effects

         There have been limited studies of TCA's effects on reproductive
    performance. TCA was found to inhibit  in vitro fertilization of
    gametes from B6D2F1 mice at a concentration of 1000 mg/litre, but was
    without effect at 100 mg/litre (Cosby & Dukelow, 1992). It is unlikely
    that these effects at very high doses relative to those that might be
    expected from human consumption of TCA at concentrations less than
    0.1% of the NOEL are of relevance in assessing risks from exposure to
    TCA in drinking-water.

    4.2.2.3  Developmental effects

         Treatment of pregnant rats with TCA at 0, 330, 800, 1200 or
    1800 mg/kg of body weight per day by gavage in a water vehicle on
    gestation days 6-15 produced dose-dependent reductions in body weight
    and length of rat pups from dams administered doses of 800 mg/kg of
    body weight per day and above. No effects were observed at 330 mg/kg
    of body weight per day. However, there was a significant and
    dose-related increase in soft tissue malformations at all doses
    studied. The mean frequency of soft tissue malformations was 3.5 
    8.7% (SD), 9.06  12.9%, 30.4  28.1%, 55.4  36.1% and 96.9  8.8% at
    0, 330, 800, 1200 and 1800 mg/kg of body weight, respectively. Most of
    the increased soft tissue abnormalities were accounted for by defects
    in the cardiovascular system. The major malformation seen was
    laevocardia. However, at doses of 800 mg/kg of body weight and above,
    a significant incidence of an interventricular septal defect was
    observed (Smith et al., 1989a).

         Saillenfait et al. (1995) found that TCA, administered to rats in
    embryo culture, began to produce consistent increases in defects at
    concentrations of 2.5 mmol/litre and above, with few or no effects
    observed at 1 mmol/litre. These defects included brain and eye
    defects, reduction in the branchial arch and otic system defects. At
    concentrations of 3.5 mmol/litre, some evidence of skeletal defects
    was observed (i.e., absence of hindlimb bud). As is discussed below,
    these concentrations can be achieved in the blood of rats administered
    high doses of TCA (>100 mg/kg of body weight) (Larson & Bull,
    1992). However, there is considerable doubt about whether these
    effects would be induced by the doses of less than 1 g/kg of body
    weight experienced by humans consuming chlorinated drinking-water. 

    4.2.2.4  Neurotoxicity

         No reports of neurotoxic effects of TCA were located.

    4.2.2.5  Toxicity in humans

         TCA is a strong acid. It is widely recognized that contact of TCA
    with the skin has the potential to produce acid burns, and ingestion
    of TCA has the potential to damage tissues of the gastrointestinal
    tract or produce systemic acidosis, even though specific studies of
    these effects do not appear in the literature. Such effects would
    occur from contact with the crystal or strong solutions of the free
    acid. However, such effects have little relevance to the production of
    low levels of TCA, as the salt, as a by-product of the chlorination of
    drinking-water.

         Indirectly, it may be presumed that TCA presents little overt
    hazard to human health because it is a major metabolite of commonly
    used solvents such as trichloroethylene and tetrachloroethylene.
    Occupational exposures to these solvents have been quite high in the
    past, but few, if any, effects of the solvents in humans have been
    attributed to TCA. Therefore, one would surmise that TCA is relatively
    non-toxic to humans under circumstances of low exposures such as those
    encountered in chlorinated drinking-water. However, these largely
    negative data do not insure against chronic hazards such as cancer or
    adverse reproductive outcomes or teratogenicity. The only reasonable
    evidence of carcinogenicity due to TCA in animals relates very
    specifically to the induction of liver tumours. If TCA's apparent mode
    of action is taken into consideration, it is difficult to identify
    other tumours that would be attributable to TCA from animal studies.
    Recent studies of workers in degreasing operations provide little
    evidence of hepatocellular tumours (Spirtas et al., 1991; Weiss,
    1996). 

    4.2.2.6  Carcinogenicity and mutagenicity

         IARC has evaluated the carcinogenicity of TCA and concluded that
    there is inadequate evidence for its carcinogenicity in humans and
    limited evidence for its carcinogenicity in experimental animals. The
    compound was assigned to Group 3: not classifiable as to its
    carcinogenicity to humans (IARC, 1995).

         TCA induces hepatocellular carcinomas when administered in
    drinking-water to male B6C3F1 mice (Herren-Freund et al., 1987; Bull
    et al., 1990; Daniel et al., 1993a). Although some of the data in the
    literature are of a preliminary nature, consistent results were
    obtained in three independent studies (Table 17). In two of these
    studies, dose-related increases in the incidence of malignant tumours
    and precancerous lesions were obtained in B6C3F1 mice at
    concentrations in water of between 1 and 5 g/litre and with as little
    as 12 months of treatment (Bull et al., 1990; Daniel et al., 1993a).
    Under similar conditions of treatment, TCA did not induce hepatic
    tumours in F344 rats (DeAngelo et al., 1997).

         In the study in mice by Pereira (1996) reported in Table 17, a
    NOAEL of 0.35 g of TCA per litre can be identified. Drinking-water
    consumption and body weight were reported only during the first
    4 weeks of the study. These data were not used; instead, it was


        Table 17. Carcinogenic effects of trichloroacetate in rodents

                                                                                                                                      
    Species        Dose        Duration   Tumour           HN & HAa                        HCb                 Reference
    (sex)          (g/litre)   (weeks)    site                                                             
                                                    Incidence   Tumour/n         Incidence    Tumour/n
                                                                (multiplicity)                (multiplicity)
                                                                                                                                      

    Mice

    B6C3F1 (M)     0           61         Liver     2/22        (0.09)           0/22         (0)              Herren-Freund et al. 
                   5           61         Liver     8/22        (0.5)            7/22         (0.5)            (1987)

    B6C3F1 (M)     0           52         Liver     1/35        0.03             0/35         0                Bull et al. (1990)
                   1           52         Liver     5/11        0.45             2/11         0.18
                   2           52         Liver     15/24       0.63             4/24         0.17
                   2           37         Liver     2/11        0.18             3/11         0.27

    B6C3F1 (M)     0           60-95      Liver     NRc         NR               6.7-10%      0.07-0.15        Daniel et al. (1993a)
                   0.05        60         Liver     NR          NR               22%          0.31
                   0.5         60         Liver     NR          NR               38%          0.55
                   4.5         95         Liver     NR          NR               87%          2.2
                   5           60         Liver     NR          NR               55%          0.97

    B6C3F1 (F)     0           52         Liver     1/40        0.03             0/40         0                Pereira (1996)
                   0.35        52         Liver     6/40        0.15             0/40         0
                   1.2         52         Liver     3/19        0.16             0/19         0
                   3.5         52         Liver     2/20        0.10             5/20         0.25
                   0           81         Liver     2/90        0.02             2/90         0.02
                   0.35        81         Liver     14/53       0.26             0/53         0
                   1.2         81         Liver     12/27       0.44             5/27         0.19
                   3.5         81         Liver     18/18       1.0              5/18         0.28
    Rats

    F344 (M)       0           104        Liver     2/23        0.09             0/23         0                DeAngelo et al. (1997)
                   0.05        104        Liver     2/24        0.08             0/24         0
                   0.5         104        Liver     5/20        0.25             0/20         0
                   5.0         104        Liver     1/22        0.045            1/22         0.045
                                                                                                                                      

    Table 17. (continued)

    a Combined hyperplastic nodule and hepatocellular adenoma.
    b Hepatocellular carcinoma.
    c NR = not reported.
    

    assumed that the average daily water consumption was about 10% of the
    animal body weight. On this basis, 0.35 g of TCA per litre is
    equivalent to approximately 40 mg/kg of body weight per day.

         The available data suggest that TCA has some tumour-promoting
    activity. Pereira (1995) and Pereira & Phelps (1996) reported that TCA
    increased the yield of both hepatocellular adenomas and hepatocellular
    carcinomas in MSU-initiated mice (Figure 4). This effect appears to be
    evident at the lowest concentration tested, 0.35 g/litre of
    drinking-water. Unlike the circumstance described above with DCA, TCA
    significantly increased the yield of hepatocellular carcinomas as well
    as hepatocellular adenomas after 362 days of treatment. An earlier
    study by Parnell et al. (1986) suggested that TCA was capable of
    promoting tumours initiated by diethylnitrosamine. However, this study
    employed gamma-glutamyl transpeptidase (GGT) as a marker for
    preneoplastic foci and was extended for only 6 months. As a
    consequence, no increase in tumour yield was noted, although there was
    a significant increase in GGT-positive foci that appeared
    dose-related. GGT is a poor marker for foci induced by peroxisome
    proliferators (Sakai et al., 1995). Therefore, this study may have
    underestimated the promoting activity of TCA.

         The mechanisms by which TCA induces tumours are not clear. TCA
    induces peroxisome proliferation in male B6C3F1 mice over the same
    dose range at which it induces hepatic tumours (DeAngelo et al.,
    1989). Unlike the situation with DCA, the induction of peroxisome
    synthesis by TCA appears to be sustained over time. Despite a large
    number of data that strongly link peroxisome proliferation with
    carcinogenesis, the actual mechanism by which such chemicals actually
    produce cancer may be only loosely associated with peroxisome
    proliferation  per se (discussed further in section 4.2.2.8).

         As noted in Table 16, the mutation spectra of mouse liver tumours
    obtained from mice treated with TCA appear to be different from those
    observed with DCA (Ferreria-Gonzalez et al., 1995). However, it is
    important to note that this result was obtained from a very limited
    number of animals (11), only five of which had hepatic tumours.
    Although none of the tumours carried the mutation that is apparently
    selected by DCA treatment, the sparseness of these data prevents a
    clear conclusion.

         Numerous mutagenicity tests have been conducted on TCA (IARC,
    1995). TCA did not induce lambda prophage in  Escherichia coli and
    was not mutagenic to  Salmonella typhimurium strains in the presence
    or absence of metabolic activation. TCA, however, reacts with dimethyl
    sulfoxide (a solvent used commonly in this assay) to form unstable
    mutagenic substances, which have not been identified (Nestmann et al.,
    1980). TCA did not induce DNA strand breaks in mammalian cells
     in vitro. Chromosomal aberrations were not induced in human
    lymphocytes exposed  in vitro to TCA neutralized to avoid the effects
    of low pH seen in cultured mammalian cells.

    FIGURE 4

         DNA strand breaks were reported in one laboratory in the livers
    of mice and rats treated 4 h previously with TCA; none was observed
    24 h after repeated daily dosing with 500 mg/kg of body weight (Nelson
    & Bull, 1988; Nelson et al., 1989). Peroxisome proliferation, as
    indicated by -oxidation of palmitoyl CoA, was observed only after
    induction of DNA damage (Nelson et al., 1989). DNA strand breakage was
    not observed in the livers of mice or rats (Chang et al., 1992). The
    reasons for the contrasting results obtained using similar techniques
    are unclear (IARC, 1995).

         Giller et al. (1997) examined the effects of TCA in the SOS
    chromotest in  Escherichia coli Pq37, the Ames fluctuation assay and
    the newt micronucleus assay. TCA was negative in the SOS chromotest
    but exhibited weak activity in the Ames fluctuation assay. Effects
    were observed at the lowest concentration, 1750 g/ml, in the absence
    of S9 fraction. This corresponds to TCA concentrations of
    approximately 10 mmol/litre in the medium. Newt larvae were found to
    have an increased frequency of micronuclei at TCA concentrations of
    80 g/ml. As with some previous studies, these tests appear to have
    been conducted with the free acid, raising issues of potential
    artefacts in the results.

         Harrington-Brock et al. (1998) studied the mutagenic activity of
    TCA in the mouse lymphoma system. A very weak positive result was
    obtained at concentrations in excess of 20 mmol/litre. Concentrations
    of TCA reached in the blood of mice treated with carcinogenic doses
    can approach the mmol/litre range, so it is possible that these
    results could be relevant to bioassay data. However, concentrations
    anticipated in drinking-water would clearly be much lower (in the low
    mol/litre range). 

         In one study, TCA induced micronuclei and chromosomal aberrations
    in bone marrow cells and abnormal sperm morphology after injection
    into Swiss mice  in vivo at doses of 125-500 mg/kg of body weight
    (Bhunya & Behera, 1987). However, Mackay et al. (1995) could not
    replicate this finding, even at doses 10-fold higher.

    4.2.2.7  Comparative pharmacokinetics and metabolism

         TCA is readily absorbed from the gastrointestinal tract in
    experimental animals and humans (Muller et al., 1974; Larson & Bull,
    1992). However, the major determinant of its blood concentrations at a
    given dose is its relatively slow clearance from blood relative to
    other HAAs. There are substantial differences in this clearance by
    different species. The half-life is 5.8 h in mice (Larson & Bull,
    1992), 9.3 h in rats (Merdink et al., 1999), 50 h in humans (Muller et
    al., 1974) and approximately 200 h in dogs (Muller et al., 1974).

         TCA is much less extensively metabolized than other HAAs found in
    drinking-water. However, one metabolite is DCA (Larson & Bull, 1992),
    which is subsequently converted to glyoxylate, glycolate and oxalate,
    as outlined in Figure 5. This in turn explains the extensive
    incorporation of radiolabel from 14C-TCA into blood (Stevens et al.,

    FIGURE 5

    1992) and tissues (Eyre et al., 1995), as glyoxylate is rapidly
    transaminated and converted to glycine.

         In mice, some of TCA's metabolism is independent of the formation
    of DCA as an intermediate. There is evidence that a significant amount
    of oxalate is formed independently of DCA formation and that some
    direct decarboxylation of trihaloacetates also occurs. The evidence
    for these pathways comes largely from studies of the metabolism of
    bromodichloroacetate (BDCA) (Xu et al., 1995). The conversion of TCA
    is, however, much slower than that of BDCA. The basis for the overall
    scheme is discussed more fully in the following section on brominated
    HAAs. The reader is also referred to Figure 5 and the accompanying
    text for explanation of the further metabolism of DCA.

    4.2.2.8  Mode of action

         The tumorigenic effects of TCA in the liver of B6C3F1 mice
    appear to be closely related to its ability to induce synthesis of
    peroxisomes and associated proteins (DeAngelo et al., 1989; Bull et
    al., 1990; Pereira, 1996). Along with a number of other peroxisome
    proliferators, TCA was shown to be capable of activating the PPAR
     in vitro at concentrations consistent with the levels that are
    achieved  in vivo (Issemann & Green, 1990). The cause-and-effect
    relationships between the activation of this receptor and the
    induction of cancer are yet to be established. Based upon marked
    increases in the numbers of peroxisomes that are observed in rodent
    species that are susceptible to this class of carcinogen and the lack
    of such responses in other mammalian species, it has been argued that
    humans are minimally sensitive to the tumorigenic effects of these
    compounds (Lake, 1995). Peroxisome proliferators are also known to
    variably affect reproduction and development. Considering the types of
    interactions in which the PPAR is involved, these observations are not
    too surprising. However, it is not clear whether a consistent pattern
    of developmental anomalies can be associated with activation of this
    receptor.

         A mouse that was genetically engineered with a targeted
    disruption PPAR-alpha gene failed to respond to the pleiotropic
    effects of peroxisome proliferators (Lee et al., 1995).

         Non-rodent species, including humans, express the PPARs, and
    peroxisome proliferator responsive elements (PPREs) have been
    identified in the promoter regions of the genes that are analogous to
    those that are induced in rodents (Varanasi et al., 1996). This
    suggests that the responsiveness to peroxisome proliferators is in
    some way modified by other factors. However, the expression of PPARs
    in humans is low.

         It is important to recognize that other mechanisms may be
    involved in TCA-induced effects. Clearly, the production of DCA as a
    metabolite could be involved in inducing effects associated with this
    metabolite. The very high rate of DCA metabolism relative to the rate
    of TCA metabolism has made it difficult to detect significant amounts
    of DCA in the blood (Merdink et al., 1998). There is little doubt that

    some DCA is formed in all species. In addition to DCA, there are a
    number of reactive and stable metabolites that could contribute to
    toxicity. At this time, there is no evidence that the acid chloride
    intermediate postulated between DCA and oxalate contributes anything
    to the toxic effects of TCA. Chronic TCA administration results in the
    deposition of lipofuscin, a sign of increased oxidative stress.
    Moreover, Austin et al. (1995, 1996) obtained evidence of increased
    lipid peroxidation and increased levels of 8-OH-dG in nuclear DNA of
    the liver of mice treated with single doses of TCA. Therefore, the
    accumulation of lipofuscin could be associated with mechanisms
    unrelated to peroxisome proliferation. The most likely
    radical-generating sources for lipid peroxidation and oxidative damage
    to DNA would be through the formation of the radical species
    (carbon-centred radicals and the peroxy radicals that would be derived
    from the reaction of molecular oxygen with the carbon-centred
    radicals). Again, there is no direct evidence to indicate that these
    processes are important to any toxicological response associated with
    TCA treatment or exposure.

         It is beyond the scope of the present review to resolve risk
    assessment issues associated with peroxisome proliferators. However,
    several points should be made specific to TCA. First, peroxisome
    proliferative responses are not genotoxic responses. Second, TCA is
    one of the weakest activators of the PPAR known (Issemann & Green,
    1990). Finally, TCA appears to be only marginally active as a
    peroxisome proliferator, even in rats (DeAngelo et al., 1989).
    Furthermore, treatment of rats with high levels of TCA in
    drinking-water does not induce liver tumours (DeAngelo et al., 1997).
    These data strongly suggest that TCA presents little carcinogenic
    hazard to humans at the low concentrations found in drinking-water. 

         The key question is whether the carcinogenic and teratogenic
    effects of TCA observed at very high doses have any relevance at the
    exposures that are obtained even under extreme conditions in
    drinking-water. However, the application of conventional uncertainty
    factors would suggest that TCA in drinking-water represents little
    hazard to humans at the concentrations normally encountered in
    chlorinated drinking-water.

    4.2.3  Brominated haloacetic acids

         Brominated HAAs are formed in waters that contain bromide, which
    strong oxidants like chlorine and ozone are capable of oxidizing to
    hypobromous acid. There are very limited data available on the
    toxicity of these chemicals. Therefore, they will be considered as a
    group. The discussion below focuses on similarities and differences
    between brominated and chlorinated HAAs.

    4.2.3.1  General toxicological properties and information on 
             dose-response in animals

         Linder et al. (1994a) found that the oral LD50 for MBA was
    177 mg/kg of body weight in adult male Sprague-Dawley rats. DBA was

    much less toxic, with an LD50 of 1737 mg/kg of body weight. The acute
    toxicities of the bromochloroacetates have not been determined.

         No studies have been published on the brominated haloacetates.
    Some of the information is available in abstract form and will be
    touched on briefly.

         Bull and co-workers (Bull & DeAngelo, 1995; Stauber et al., 1998)
    conducted a series of experiments with DBA, bromochloroacetate (BCA)
    and BDCA administered to male B6C3F1 mice at concentrations ranging
    from 0.2 to 3 g/litre in their drinking-water. The data obtained from
    these studies suggest that toxicological effects are observed in
    approximately the same concentration ranges as for DCA- and
    TCA-induced toxic effects.

         The principal target organ in mice was identified as the liver
    for all three brominated HAAs, but the nature of the effects on the
    liver appears to be somewhat different for each compound. All the
    brominated HAAs produce hepatomegaly, but glycogen accumulation and
    cytomegaly are more prominent with BCA and BDCA than with DBA.
    Conversely, DBA was reported to produce increases in
    cyanide-insensitive acyl CoA activity in the liver. BCA and BDCA
    produced only small and inconsistent effects on this marker of
    peroxisome proliferation. Thus, it appears that DBA shares TCA's
    ability to induce peroxisome proliferation, whereas BCA and BDCA
    appear to produce effects much more like those induced by DCA.
    Consistent with this, the severity of the hepatomegaly was BCA > BDCA
    > DBA.

         Administration of single doses of BDCA, BCA and DBA by gavage in
    water as the vehicle induced increases in thiobarbituric acid reactive
    substances (TBARS) and increased the 8-OH-dG content of nuclear DNA in
    the liver of male B6C3F1 mice at doses as low as 30 mg/kg of body
    weight (Austin et al., 1996). These effects were significantly greater
    than those observed with TCA and DCA and tended to increase as the
    bromine substitution increased within each series. The order of
    potency was DBA = BCA > BDCA > DCA > TCA. Increases in 8-OH-dG
    content were more rapid and more sustained with the brominated HAAs.
    This indicates that brominated HAAs do induce oxidative stress.

         Dose-related increases in the 8-OH-dG content of nuclear DNA of
    the liver were observed when DBA and BCA were administered in
    drinking-water to male B6C3F1 mice for periods from 3 to 10 weeks at
    concentrations of 0.5 g/litre and above (Parrish et al., 1996). The
    effect of BDCA in drinking-water was not evaluated. It is noted that
    these effects were observed in the same range in which liver tumours
    are induced in mice (discussed further below).

    4.2.3.2  Reproductive effects

         MBA and DBA have been examined for spermatotoxic effects in rats.
    MBA did not affect parameters related to male reproductive function
    (Linder et al., 1994a) when administered to rats as a single dose of
    100 mg/kg of body weight or at 25 mg/kg of body weight per day

    administered repeatedly for 14 consecutive days. In contrast, DBA
    produced degenerating, misshapen epididymal sperm and abnormal
    retention of step 19 spermatids following single doses of DBA in the
    range 1000-2000 mg/kg of body weight. Caput sperm counts were
    significantly reduced on the second day, and substantial reductions of
    cauda sperm counts were observed at 14 and 28 days after treatment.
    Serum testosterone levels were significantly depressed on day 2 but
    returned to control levels by day 14. Sperm displayed defects in
    development of the shape of sperm heads. Progressive motility of sperm
    was significantly reduced at 14 and 28 days after treatment.

         A subsequent paper by the same group described a study in which
    doses of 0, 10, 30, 90 or 270 mg of DBA per kg of body weight were
    administered to rats for 14 consecutive days (Linder et al., 1994b).
    Marked effects on epididymal sperm counts and sperm morphology effects
    were observed at the highest dose. Approximately 5% of the caput sperm
    were fused. In contrast to the single-dose studies, serum testosterone
    levels appeared to be unaffected. Spermiation also appeared to be
    mildly affected, with step 19 spermatids being retained beyond stage
    VIII in animals dosed with as little as 10 mg/kg of body weight per
    day.

         DBA was administered to rats for up to 79 days at 0, 2, 10, 50 or
    250 mg/kg of body weight per day by gavage in water. Male fertility
    was compromised during the second week of treatment at the high dose.
    This effect appeared to result from behavioural changes, because
    artificial insemination with sperm collected on day 9 of treatment
    produced offspring. By day 15, however, no offspring were produced
    with either biological or artificial insemination, indicating that
    significant qualitative alterations had occurred in the sperm. Indeed,
    the 50 mg/kg of body weight per day dose produced abnormal morphology,
    decreased motility and decreases in epididymal sperm counts. However,
    rats treated at this dose remained fertile. While no effects on sperm
    quality were observed at lower doses, reproductive performance
    appeared depressed at doses as low as 10 mg/kg of body weight per day,
    suggesting that these lower doses modified behaviour (Linder et al.,
    1995).

         Histopathological changes were observed in the testis and
    epididymis of rats gavaged daily for 2-79 days with DBA. On treatment
    day 2, abnormal retention of step 19 spermatids was observed in
    animals given the highest dose of 250 mg/kg of body weight per day.
    Additional changes on day 5 included the fusion of mature spermatids
    and the presence of atypical residual bodies (ARB) in the epithelium
    and lumen of stage X-XII seminiferous tubules. By day 9, ARB were seen
    in most stages of the seminiferous epithelial cycle and in the caput
    epididymis. On day 16, distorted sperm heads were recognized in step
    12 and older spermatids, and luminal cytoplasmic debris was found
    throughout the epididymis. On day 31, there was vacuolation of the
    Sertoli cell cytoplasm, extensive retention of step 19 spermatids near
    the lumen of stage IX and X tubules, and vesiculation of the acrosomes
    of late spermatids. Marked atrophy of the seminiferous tubules was
    present 6 months after 42 doses of 250 mg/kg of body weight per day.
    ARB and retention of step 19 spermatids were observed after 31 and 79

    doses of 50 mg/kg of body weight per day, and increased retention of
    step 19 spermatids was seen in several rats dosed with 10 mg/kg of
    body weight per day. No abnormalities were detected at 2 mg/kg of body
    weight per day. The changes suggest that the testicular effects of DBA
    are sequelae to structural or functional changes in the Sertoli cell
    (Linder et al., 1997b).

    4.2.3.3  Neurotoxicity

         Reference was made to neurotoxic effects in male Sprague-Dawley
    rats treated with DBA in reproductive studies (Linder et al., 1994a,b,
    1995). However, no specific investigations of the neurotoxic effects
    of brominated HAAs appear to be available.

    4.2.3.4  Toxicity in humans

         There are no studies on the health effects of brominated HAAs in
    humans.

    4.2.3.5  Carcinogenicity and mutagenicity

         No formal reports have been made of carcinogenic or mutagenic
    effects of brominated HAAs. Indications that DBA, BCA and BDCA share
    the hepatocarcinogenic effects of TCA and DCA in B6C3F1 mice have
    been referred to in an abstract (Bull & DeAngelo, 1995).

         Giller et al. (1997) examined the genotoxicity of MBA, DBA, and
    tribromoacetic acid (TBA) in the SOS chromotest in  Escherichia coli 
    PQ37, the Ames fluctuation assay utilizing  Salmonella typhimurium 
    strain TA100 and the newt micronucleus assay in larvae at stage 53 of
    the developmental table. MBA was negative in the SOS chromotest at
    levels as high as 1000 g/ml and in the newt micronucleus assay.
    However, it was active at concentrations as low as 20 g/ml in the
    Ames fluctuation assay with S9 fraction added to the incubation
    medium. DBA was positive with and without S9 fraction in the SOS
    chromotest, requiring 100 g/ml in the former case and 200 g/ml in
    the latter. Thus, it is about 5 times as potent as DCA in this test.
    In the Ames fluctuation assay, DBA was active at 10 g/ml without S9
    and at 30 g/ml with S9. TBA was active in the SOS chromotest at
    100 g/ml with S9 but required 2000 g/ml for activity in the Ames
    fluctuation assay without S9.

    4.2.3.6  Comparative pharmacokinetics and metabolism

         Metabolism of brominated HAAs has received little attention. Xu
    et al. (1995) studied the metabolism of BDCA in male B6C3F1 mice. As
    predicted, substitution of a bromine for a chlorine in TCA resulted in
    a substantially greater extent of trihaloacetate metabolism. Whereas
    45% of a 100 mg/kg of body weight dose of TCA was eliminated unchanged
    in the urine of mice within 24 h, less than 4% of the same dose of
    BDCA was found in the urine. At lower doses, only a fraction of a
    percent of the BDCA was eliminated unchanged.

         There is evidence for substantial conversion of BDCA to DCA in
    both rats and mice (Xu et al., 1995; Schultz et al., 1999). Because of
    the established ability of DCA to induce hepatic tumours in both mice
    and rats (discussed above), this may have implications for assessing
    the risk associated with this compound (see Figures 3 and 5 for a
    general description of dihaloacetate and trihaloacetate metabolism,
    respectively). 

         The metabolism of BDCA is differentially modified in mice and
    rats as doses are increased. Xu et al. (1995) found that the kinetics
    of carbon dioxide production from 1-14C-BDCA suggested an efficient
    conversion of BDCA to carbon dioxide through DCA at low doses, but a
    direct decarboxylation reaction became important as doses approached
    100 mg/kg of body weight (Xu et al., 1995; Austin & Bull, 1997). This
    complex activity was not observed in rats (Schultz et al., 1999), in
    that a progressively smaller fraction of the dose is converted to
    carbon dioxide as dose is increased. This suggests that direct
    decarboxylation plays a less important role in the metabolism of BDCA
    in rats than in mice.

         The ratios of urinary metabolites produced by mice and rats
    suggest that there are some substantive differences in the metabolism
    of BDCA in the two species. Mice (Xu et al., 1995) produce much higher
    amounts of oxalate (about 30% of the orally administered dose) than do
    rats (about 20%) (Schultz et al., 1999). The much greater conversion
    of BDCA to oxalate than for equivalent doses of DCA suggests that much
    of the extra oxalate seen in mouse urine arises from reductive
    dehalogenation of BDCA, followed by peroxy radical formation and
    decomposition to oxalate (Xu et al., 1995).

         As the BDCA dose is increased from 20 to 100 mg/kg of body weight
    in the rat, the fraction of the dose that is eliminated in the urine
    as DCA increases from about 2% to 13% (Schultz et al., 1999), whereas
    in mice the increase is from 0.2% to approximately 3%. In the rat,
    increasing the dose from 5 mg/kg of body weight to 20 or 100 mg/kg of
    body weight was associated with a significantly extended half-life of
    both BDCA and DCA (from 0.9 to 3.7 h). The increased half-life appears
    to be attributed to saturation of the conversion of BDCA to DCA.
    Despite this inhibition of DCA formation, the blood levels of DCA
    observed are actually higher than those observed when equivalent doses
    of DCA itself are administered. Thus, high doses of BDCA would appear
    to inhibit the metabolic clearance of DCA as well. Consequently, the
    toxicology of BDCA should share many of the attributes of DCA
    toxicology in the rat. Comparable data do not exist on the blood
    levels of DCA achieved by BDCA administration in mice, but the urinary
    levels of DCA metabolites appear to be largely offset by the
    conversion of BDCA to oxalate, suggesting that the blood levels of DCA
    would be significantly lower in mice than in rats.

         In animals previously treated with 1 g of DCA per litre in
    drinking-water, metabolism of BDCA to carbon dioxide is significantly
    increased in mice (Austin & Bull, 1997). This appears to be associated
    with an increase in the capacity for the direct decarboxylation of

    BDCA to form BDCM and carbon dioxide. Therefore, some effects of BDCA
    in mice may be attributed to the formation of BDCM.

         These data and data indicating that pretreatment with DCA also
    affects its own metabolism suggest that there may be significant
    interactions in the toxicity of these chemicals at high doses. It
    remains to be seen if such interactions occur at treatment
    concentrations that more closely approximate those observed in
    drinking-water.

    4.2.3.7  Mode of action

         It is premature to attempt a definitive discussion of the
    mechanisms by which brominated HAAs induce tumours. There are
    suggestions of mechanisms that need to be pursued with this class, but
    there are really no data that demonstrate how they contribute to the
    induction of cancer, nor is there a firm basis for considering whether
    these mechanisms are relevant at the low concentrations of these
    chemicals found in drinking-water.

         The mechanisms associated with the carcinogenic effects of HAAs
    include those identified for DCA and TCA. It is apparent that more
    than one mechanism is responsible for the effects of this class and
    that the importance of these mechanisms to the activity of individual
    members of the class varies. In part, these differences in mechanism
    can be related to the differences in tumour phenotypes that are
    induced. One phenotype seems to be associated with prior
    characterizations of tumours induced by peroxisome proliferators and
    is induced by TCA (DeAngelo et al., 1989; Stauber et al., 1998). The
    second phenotype involves glycogen-poor tumours that stain heavily
    with antibodies to c-Jun and c-Fos. This phenotype is produced by DCA.
    These effects are probably produced by selection of lesions with
    differing defects in cell signalling pathways that control the
    processes of cell division and cell death.

         Based upon the data of Giller at al. (1997), the brominated HAAs
    are about 10-fold more potent than their chlorinated analogues in
    their ability to induce point mutations. This does not establish that
    they are inducing cancer by mutagenic mechanisms  in vivo, but this
    activity will have to be taken into account as data on their
    carcinogenic activity become more complete.

         The HAAs vary widely in their ability to induce oxidative stress
    and to elevate the 8-OH-dG content of nuclear DNA of the liver. This
    property becomes increasingly apparent with the brominated compounds
    (Austin et al., 1996; Parrish et al., 1996). It is notable that the
    brominated analogues are not more potent inducers of hepatic tumours
    than the corresponding chlorinated HAAs. Therefore, it is doubtful
    that this mechanism is the most important determinant of this effect.

         No specific mechanisms have been associated with the effects of
    DBA on male reproduction or as a developmental toxin. However, it
    would be surprising if the effects on cell signalling systems that

    appear to be involved in the carcinogenic responses do not also
    contribute to these effects. 

    4.2.4  Higher molecular weight halogenated acids

         Studies of by-product formation from humic and fulvic acids in
    the 1970s and early 1980s demonstrated that there is a complex array
    of halogenated carboxylic acids in addition to the HAAs (Christman et
    al., 1983). Some of these acids have been identified in drinking-water
    as well, but, as discussed elsewhere in this document, both the scope
    and quantitative nature of the data that are available for
    drinking-water itself are limited. Little attention has been paid to
    these higher molecular weight acids in either the toxicological or
    epidemiological literature. However, it is important to recognize that
    they make up the bulk of the total organic halogen that is present in
    drinking-water.

         Beyond the HAAs, the most studied group of halogenated acids are
    the chloropropionic acids. As with DCA, the major impetus for these
    studies has been potential therapeutic applications rather than
    concerns over exposure to the compounds as contaminants of
    drinking-water. Therefore, there are few data that have been developed
    for hazard identification purposes, and even less information is
    available on dose-response on effects other than those being explored
    for therapeutic purposes.

         By far the most data on chloropropionic acids exist for
    2-chloropropionate (2-CP). This compound shares the hypoglycaemic
    effects of DCA. It was first used as an experimental tool to segregate
    direct effects of DCA from those of its metabolites. Oxalate and
    glyoxylate are two metabolites of DCA that seem to be responsible for
    its effects on gluconeogenesis from lactate (Crabb & Harris, 1979).
    Since 2-CP is not metabolized to these compounds and failed to inhibit
    gluconeogenesis, these data effectively argue that this effect of DCA
    is largely attributable to these two metabolites. In contrast, 2-CP
    still decreased blood glucose in rats when infused intravenously at a
    rate of 300 mg/kg of body weight per hour. As with DCA, 2-CP increased
    concentrations of circulating ketone bodies but significantly reduced
    blood concentrations of lactate. As had been previously demonstrated
    with DCA, concentrations of 1 and 5 mmol of 2-CP per litre
    significantly enhanced the activity of pyruvate dehydrogenase.

         Yount et al. (1982) compared the effects of DCA and 2-CP in mice
    and rats. They extended these investigations into the area of
    comparative toxicities of the two compounds as well. The acute oral
    LD50 for 2-CP in fasted ICR mice was 15.4  0.1 mmol/kg of body
    weight, making it approximately twice as toxic as DCA on a molar
    basis. However, neither compound is really very toxic acutely (2-CP =
    1671 mg/kg of body weight; DCA = 4100 mg/kg of body weight).

         Male Wistar rats were administered the sodium salts of 2-CP and
    DCA at 0.04 mol/kg of feed for 12 weeks. These concentrations
    correspond to about 4.3 and 5.1 g of 2-CP and DCA per kg of feed,
    respectively. Actual doses to the animals are difficult to calculate

    accurately because the body weights were not provided and because
    significant effects of both compounds on body weight gain were noted.
    If an average weight of 300 g is assumed for rats for the duration of
    the experiment (i.e., study was started with weanling rats), the doses
    can be estimated to be roughly 300 mg/kg of body weight per day for
    DCA and 250 mg/kg of body weight per day for 2-CP.

         Both DCA and 2-CP decreased the growth rate and food consumption
    of treated rats and caused neurotoxic effects (e.g., hind limb
    weakness). 2-CP treatment caused testicular abnormalities and
    significantly lowered plasma triacylglycerol levels compared with
    control or DCA-treated rats. In mature rats, total serum ketone bodies
    were increased by DCA but not by 2-CP (Yount et al., 1982). 

         2-CP, as either the L- or D-isomer, is rapidly and extensively
    metabolized in the liver cytosol by a mechanism that depletes GSH
    (Wyatt et al., 1996). Significant depletion of non-protein sulfhydryl
    content (primarily GSH) was observed with single acute oral doses of
    62.5 mg/kg of body weight and above in male Alderley Park
    Wistar-derived rats. This effect was observed to be maximal at 4 h and
    returned to control values in approximately 48 h. A slower depletion
    of non-protein sulfhydryl content was observed in the cerebellum and
    forebrain with doses of 750 mg/kg of body weight. The depletion was
    observed to be maximal about 24 h after administration in the
    cerebellum and between 12 and 24 h in the forebrain. The depletion in
    the cerebellum occurred with doses that also resulted in the induction
    of granule cell necrosis in the cerebellum. Although depletion of
    non-protein sulfhydryl groups tended to recover in the liver, the
    effect in the cerebellum appeared to be partially cumulative with each
    daily dose. 

         As with DCA, the largest capacity for 2-CP metabolism appears to
    be in the cytosolic fraction of the hepatocyte, whereas metabolism in
    microsomes is quite small. L-CP appears to be slightly more active in
    depleting cytosolic fractions of GSH. In the process, the metabolite
    2- S-glutathionylpropanic acid is formed. On the basis of the
    stoichiometry between GSH depletion and formation of this product and
    prior observations of Polhuijs et al. (1989, 1991) with
    2-bromocarboxylic acids, it was concluded that 2-CP was a substrate
    for a theta-class GST.

    4.3  Haloaldehydes and haloketones

         A diverse set of halogenated aldehydes and ketones are formed in
    the disinfection of drinking-water. The most important in terms of
    having been identified in drinking-water is trichloroacetaldehyde or
    chloral hydrate, which is discussed separately. The remainder of the
    group is discussed collectively. From the perspective of
    drinking-water problems, the class has received very little attention,
    but its members have been identified as key metabolites of chemicals
    such as trichloroethylene, vinyl chloride and dibromochloropropane
    (Omichinski et al., 1988; Spengler & Singer, 1988). The principal
    evidence that they are formed comes from studies of chlorinated humic
    and fulvic acids (Meier et al., 1983, 1985a) and of kraft pulp

    chlorination (Kringstad et al., 1981). However, when specifically
    looked for in drinking-water using techniques with adequate analytical
    sensitivity, selected members of the group have been found in
    concentrations that are somewhere between those of the HANs and HAAs
    (Coleman et al., 1984).

         These two classes of chemicals played an important role in the
    initial studies of chemicals formed in the chlorination of
    drinking-water. The interest in them was sparked by the fact that they
    were among the first by-products to be identified that contributed to
    the mutagenic activity produced in chlorinated water (Cheh et al.,
    1980). Attention faded from this group when MX was identified, because
    this compound contributed the major proportion of the mutagenic
    activity produced by chlorine (Meier, 1988). Because these chemicals
    represent an additional group of mutagenic chemicals, some of which
    are capable of initiating skin tumours, it is important to acknowledge
    their presence. However, because of the lack of data that can be used
    to assess the degree of hazard, the present review will not be
    comprehensive.

    4.3.1  Chloral hydrate (trichloroacetaldehyde, chloral)

         Trichloroacetaldehyde (chloral) is hydrated in water and in the
    body to form the well-known sedative-hypnotic, chloral hydrate. Most
    toxicological and metabolic studies have been conducted with chloral
    hydrate rather than trying to deal with maintaining the aldehyde in
    the dehydrated state.

         It should be recognized that where there are significant amounts
    of bromide in treated drinking-water, the use of oxidants will also
    produce brominated analogues of chloral hydrate. This has been poorly
    documented with the trihaloacetaldehydes, but the brominated analogues
    have been observed in a system where chlorination reactions with
    fulvic acid are conducted in the presence of high bromide
    concentrations (Xie & Reckhow, 1993). The lack of data from actual
    water supplies is in large part due to the lack of appropriate
    analytical standards. In a broad sense, however, the ratios of chloral
    hydrate, bromodichloroacetaldehyde, dibromochloroacetaldehyde and
    tribromoacetaldehyde should more or less parallel those seen with the
    analogous trihaloacetates. A second problem is that there is not a
    significant body of toxicological literature available for these
    analogues because they are not utilized in commerce. This represents a
    significant data gap because bromine substitution can be anticipated
    to result in greater metabolism of the trihaloacetaldehydes, as has
    been demonstrated for the trihaloacetates (Xu et al., 1995). As a
    consequence, an exaggeration of those effects of chloral hydrate that
    are secondary to metabolism through reactive intermediates might be
    expected (Hoz et al., 1991). Owing to the lack of data, however, it is
    inappropriate to speculate much further. Therefore, the remainder of
    this section will specifically address the data that are available for
    chloral hydrate.

    4.3.1.1  General toxicological properties and information on 
             dose-response in animals

         Chloral hydrate is primarily known for its depressant effects on
    the central nervous system (Gilman et al., 1991). The usual doses
    required to produce central nervous system depression in humans range
    from about 500 to 2000 mg in adults. These effects do not appear to
    have been extensively studied in experimental animals. 

         Chloral hydrate was administered for 90 days in drinking-water to
    male and female Sprague-Dawley rats at concentrations of 300, 600,
    1200 or 2400 mg/litre. Hepatocellular necrosis was observed in 2 of
    10 male rats treated with concentrations of either 1200 or 2400 mg of
    chloral hydrate per litre (Daniel et al., 1992b). No liver damage was
    seen in female rats. The necrosis observed at the highest dose was
    indicated as being more severe than that observed at 1200 mg/litre,
    providing some indication of a dose-response. It is of interest that
    there was no sign of hepatomegaly produced by chloral hydrate at
    either this dose or any lower dose. If it is assumed that rats drink
    about 10% of their body weight per day, the level of 1200 mg/litre
    corresponds to approximately 120 mg/kg of body weight per day, or
    about 8 g for a 70-kg human, close to the estimated doses in the study
    by van Heijst et al. (1977).

         In contrast to findings in rats, male CD-1 mice displayed
    hepatomegaly when doses of 144 mg of chloral hydrate per kg of body
    weight per day were administered by gavage for a period of 14 days
    (Sanders et al., 1982). No effect was observed at 14.4 mg/kg of body
    weight per day. Other organs remained normal at gross necropsy, and
    there were no signs of altered serum enzyme levels (e.g., LDH or serum
    glutamate-pyruvate transaminase [SGPT]) or altered BUN. These data
    suggest that chloral hydrate is not cytotoxic at these doses. This
    short-term experiment was followed by a second experiment in which
    mice (140 per sex per group) were administered 70 or 700 mg of chloral
    hydrate per litre in drinking-water for up to 90 days. These levels
    were estimated to yield the same doses as those used in the 14-day
    range-finding study and averaged 18 or 173 mg/kg of body weight per
    day for female mice and 16 or 160 mg/kg of body weight per day for
    male mice. Male mice displayed hepatomegaly after 90 days of treatment
    at both the low and high doses (Sanders et al., 1982). Small, but
    statistically significant, increases in LDH and serum
    glutamate-oxaloacetate transaminase (SGOT)were observed at the high
    dose, but not at 70 mg/litre. The female mice did not demonstrate the
    hepatomegaly observed in males, but they did show alterations in
    hepatic microsomal parameters. These data are of potential
    significance in considering the greater sensitivity of mice (albeit of
    another strain) to the hepatocarcinogenic effects of chloral hydrate
    as compared with rats.

         Chloral hydrate has the potential of interacting with other drugs
    by direct and indirect means. Most commonly cited is a potentiation
    with alcohol that has been associated with the so-called "Mickey Finn"
    (Gilman et al., 1991). This is attributed to interactions with common
    steps in the metabolism of both chloral hydrate and alcohol (Gessner &

    Cabana, 1970; Sellers et al., 1972). Combined exposure appears to
    interfere with glucuronidation of the metabolite trichloroethanol and
    with its conversion to TCA (Kaplan et al., 1967; Sellers et al.,
    1972). Moreover, trichloroethanol is probably the metabolite of
    chloral hydrate that is primarily responsible for its central nervous
    system depressant effects (Gilman et al., 1991); thus, there is
    undoubtedly somewhat of a pharmacological interaction as well.
    Increased sensitivity to the hypnotic activity of chloral hydrate in
    paraoxon-treated animals has been suggested to result from an
    increased sensitivity of the brain to hypoxia (Koepke et al., 1974).
    The other major interaction associated with chloral hydrate is the
    ability of its metabolites, particularly TCA, to compete for binding
    sites on plasma proteins. This type of interaction has been held
    responsible for increased pharmacological and toxicological reactions
    to warfarin (Sellers & Koch-Weser, 1970; Koch-Weser et al., 1971) and
    bis-hydroxycoumarin (Cucinell et al., 1966).

    4.3.1.2  Toxicity in humans

         The primary effect seen with ingestion of chloral hydrate is
    central nervous system depression, the basis of its use in
    therapeutics. The usual dose recommended for sedation is 250 mg 3
    times daily. The hypnotic dose is generally given as 500-1000 mg, but
    2000 mg is required to be effective in many adults (Gilman et al.,
    1991). Neonates are frequently treated with doses of chloral hydrate
    in the range of 30-40 mg/kg of body weight (Lambert et al., 1990). In
    recent years, chloral hydrate has become popular for sedating
    subjects, particularly children, to aid in performing diagnostic
    procedures such as CT scans, electroencephalograms (EEG) and
    electrocardiograms (ECG), where relatively higher doses have been
    utilized (32-80 mg/kg of body weight, Silver & Steir, 1971; 100 mg/kg
    of body weight, Farber & Abramow, 1985).

         The most important acute toxic effect is the production of
    cardiac arrhythmias. Where doses in adults have been estimated, they
    are considerably above those commonly used for therapeutic purposes,
    most often in excess of 8 g. Most of the available studies involved
    poisoning or overdose situations. However, one study closely examined
    the induction of arrhythmias in paediatric cases undergoing EEGs
    (Silver & Steir, 1971). The dose range was 32-80 mg/kg of body weight,
    and in only 2 of 12 subjects was a sinus arrhythmia associated with
    the administration of chloral hydrate. This suggests that doses at the
    lower end of this range may approximate a threshold. Doses in the
    range of 96 mg/kg of body weight and above have consistently been
    shown to produce arrhythmias in children (Nordenberg et al., 1971;
    Farber & Abramow, 1985; Hirsch & Zauder, 1986). In a 70-kg adult, 32
    mg/kg of body weight would equal a dose of 2240 mg.

         Lower doses of chloral hydrate have been associated with some
    adverse side-effects. A study of newborns administered chloral hydrate
    indicated a high incidence of direct hyperbilirubinaemia (Lambert et
    al., 1990). This effect was associated more with continuous use than
    with an acute dose, however. The dose rate to affected (40 mg/kg of
    body weight) and unaffected (33 mg/kg of body weight) neonates was not

    significantly different. The total dose in affected children was
    1035 mg relative to 135 mg in the unaffected children. Thus, a more
    protracted use of chloral hydrate in the affected group was apparently
    responsible for the hyperbilirubinaemia observed. These data suggest
    that there is little concern over this effect with single doses of
    chloral hydrate. Because neonates are generally thought to be more
    sensitive to hyperbilirubinaemia, this effect is probably of less
    concern for adults.

         There have been occasional reports of liver damage induced by
    high doses of chloral hydrate by ingestion (van Heijst et al., 1977;
    Gilman et al., 1991). The short-term data from mice, discussed
    previously, support the conclusion that an effect on the liver is
    unlikely to be observed in humans until very high doses are reached.
    Longer-term exposures (e.g., months) lead to some enlargement of the
    liver, if humans are as sensitive as mice in this regard, but the
    doses remain considerably above the fraction of a g/kg of body weight
    that would be expected from most chlorinated drinking-waters. The
    clinical literature suggests that humans are closer to rats in the
    sensitivity of their livers to chloral hydrate.

         A single modern case of fixed cutaneous eruptions was noted in
    the literature (Miller et al., 1966). This was associated with a
    therapeutic dose of chloral hydrate (500 mg) in a 57-year-old man.
    These lesions are termed "fixed" because they tend to occur at the
    same locations on the body with repeated exposures. An earlier case
    was reported in 1878. This seems to be a rare side-effect of chloral
    hydrate that is completely reversible. There have been other reported
    skin reactions to chloral hydrate ingestion, but these appear to be
    relatively rare as well (Almeyda & Levantine, 1972).

    4.3.1.3  Carcinogenicity and mutagenicity

         IARC has evaluated the carcinogenicity of chloral and chloral
    hydrate and concluded that there is inadequate evidence for their
    carcinogenicity in humans, limited evidence for the carcinogenicity of
    chloral hydrate in experimental animals and inadequate evidence for
    the carcinogenicity of chloral in experimental animals. Both chloral
    and chloral hydrate were assigned to Group 3: the compounds are not
    classifiable as to their carcinogenicity to humans (IARC, 1995).

         A more recent concern with chloral hydrate has been findings that
    it has some genotoxic and carcinogenic effects in animals and  in 
     vitro test systems (Table 18). Interpretation of some of these data
    is difficult, as many investigators failed to document the purity of
    the chemical tested. Nevertheless, chloral hydrate tends to be
    positive in  Salmonella typhimurium strain TA100, but not in TA98
    (Waskell, 1978) or TA1535 (Bignami et al., 1980). The activity towards
    TA100 was very weak in one assay (Waskell, 1978) and substantially
    greater in another (Bignami et al., 1980). It was notable that Waskell
    (1978) recrystallized chloral hydrate from alcohol 6 times before
    subjecting it to test. Bignami et al. (1980) also found that chloral
    hydrate was capable of inducing point mutations in other test systems.
    A third group found chloral hydrate to be negative in TA98, TA100,

        Table 18. Results of genotoxicity assays of chloral hydrate
                                                                                                  
    Dose             Test system        Result                                Reference
    [conc.]a
                                                                                                  
    10 mg/plate      Salmonella         Chloral hydrate purified before use   Waskell (1978)
                     TA100              Positive, S9 enhanced slightly
                     TA98               Negative

    1-5 mg/plate     Salmonella         Negative                              Bignami et al. 
                     TA1535             Positive; decreased with S9           (1980)
                     TA100

    2-10 mg/plate    Streptomyces       Positive                              Bignami et al. 
                     coelicolor                                               (1980)

    1-10 mg/plate    Aspergillus        Positive                              Bignami et al. 
                     nidulans                                                 (1980)

    82.7-413.5       Mouse              Produces non-disjunction              Russo et al. 
    mg/kg bw, i.p.   spermatocytes      Increases hyperhaploidy at all        (1984)
                                        doses tested

    [5-20            Saccharomyces      Increased mitotic gene conversion     Bronzetti et al. 
    mmol/litre]      cerevisiae         at trp+ locus in presence of S9;      (1984)
                                        ilv+ revertants not affected

    [5-10            Aspergillus        Increased mitotic segregation at      Crebelli et al. 
    mmol/litre]      nidulans           both doses, purity stated 99%         (1985)

    [1-25            Saccharomyces      Inhibits sporulation and increased    Sora & Carbone 
    mmol/litre]      cerevisiae         diploid and disomic clones            (1987)

    [25-250          DNA-protein        Negative                              Keller & Heck 
    mmol/litre]      cross-links in                                           (1988)
                     isolated rat 
                     liver nuclei

    [0.001-0.003%]   Chinese hamster    Increased number of aneuploid cells   Furnus et al. 
                     embryonic          at all doses; other chromosomal       (1990)
                     diploid cells      aberrations produced at two higher 
                                        concentrations

    0.008-5          Salmonella         All negative, purity of compound      Leuschner & Leuschner 
    mg/plate         TA98               specified                             (1991)
                     TA100
                     TA1535
                     TA1537
                     TA1538

    500 mg/kg        Mouse              Negative                              Leuschner & Leuschner 
    bw, i.p.         micronucleus                                             (1991)
                                                                                                  

    Table 18. (continued)

                                                                                                  
    Dose             Test system        Result                                Reference
    [conc.]a
                                                                                                  
    100-1000         Chromosome,        Negative at 6 and 24 h                Leuschner & Leuschner 
    mg/kg bw,        rat bone                                                 (1991)
    p.o.             marrow

    250-750          Human peripheral   Increased hyperdiploid nuclei, and    Vagnarelli et al. 
    g/ml            blood              percentage of aneuploid mitosis;      (1990)
                     lymphocytes        purity specified at 99%

    300 g/ml        Ames               Metabolism-dependent positive         Giller et al. 
                     fluctuation        response in TA100; negative at        (1995)
                     assay              100 g/ml

    200 g/ml        Newt larvae        Micronuclei in peripheral blood       Giller et al. 
                     micronucleus       erythrocytes in vivo; negative at     (1995)
                     assay              100 g/ml
                                                                                                  

    a  Abbreviations used: conc. = concentration; bw = body weight; i.p. = intraperitoneal; 
       p.o. = per os.
    
    TA1535, TA1537 and TA1538 (Leuschner & Leuschner, 1991), and the
    purity of the chloral hydrate was specified. Keller & Heck (1988)
    could find no evidence of DNA-protein cross-links with chloral hydrate
    treatment of isolated rat liver nuclei. A number of laboratories have
    shown that chloral hydrate is capable of producing chromosomal
    aberrations  in vitro (Bronzetti et al., 1984; Crebelli et al., 1985;
    Sora & Carbone, 1987; Furnus et al., 1990; Vagnarelli et al., 1990),
    including aneuploid cells. Chromosomal effects appear to be more
    consistently observed, and, thus, these results are more convincing.
    Even in these cases, however, the purity of the compound tested was
    generally not determined.

         Chloral hydrate has been extensively studied as a potentially
    genotoxic agent. It has been evaluated in the recommended screening
    battery and several other assays, including genetic alterations in
    rodent germ cells. Chloral hydrate is positive in bacterial mutation
    tests, indicating that it is capable of inducing point mutations
    (Waskell, 1978; Haworth et al., 1983; Giller et al., 1995). It is
    positive in the mouse lymphoma assay for mutations at the  Tk locus
    (Harrington-Brock et al., 1998). Chloral hydrate is also positive in
    several other  in vitro assays for genetic damage. It induces
    anueploidy in Chinese hamster embryonic fibroblasts (Natarajan, 1993),
    Chinese hamster pulmonary lines LUC2 and Don.Wq.3H (Warr et al., 1993)
    and human peripheral blood lymphocytes (Sbrana et al., 1993). Positive
    micronuclei induction was observed in Chinese hamster cells (Lynch &
    Parry, 1993) and human peripheral blood lymphocytes (Ferguson et al.,
    1993), and chromosomal aberrations were found in Chinese hamster

    embryonic diploid cells (Furnus et al., 1990). It is not clear whether
    chloral hydrate is capable of inducing genetic damage  in vivo. There
    is a mixture of positive and negative  in vivo data. Russo & Levis
    (1992) found chloral hydrate to be capable of inducing aneuploidy in
    mouse spermatocytes. Two different groups observed an increase in
    micronuclei in mouse spermatids when treatment involved exposure of
    spermatogonia stem cells (Allen et al., 1994; Nutley et al., 1996).
    Russo et al. (1992) found chloral hydrate to induce micronuclei in
    mouse bone marrow erythrocytes. Other laboratories have found chloral
    hydrate to be negative in  in vivo experiments (Xu & Adler, 1990;
    Adler, 1993). So far, chloral hydrate has been found to give negative
    test results in studies with mouse oocytes (Mailhes et al., 1993).
    Although chloral hydrate can induce a variety of genetic events
    (mutation, aneuploidy, structural chromosomal aberrations), it does so
    with a very low potency.

         Chloral hydrate has been reported to produce hepatic tumours in
    male B6C3F1 mice in two studies. One study administered by gavage a
    single dose of 5 or 10 mg of chloral hydrate per kg of body weight to
    groups of 25 and 20 male mice at 15 days of age (Rijhsinghani et al.,
    1986). The response to the 10 mg/kg of body weight dose led to
    statistically elevated levels of tumours between 48 and 92 weeks, but
    the results are based on the appearance of three adenomas and three
    carcinomas among eight animals. Moreover, the historical control
    incidence of hepatic tumours in the male of this hybrid is generally
    about 25% and has been reported to be in excess of 40% in individual
    studies. Thus, the small numbers of animals make it difficult to give
    much credence to the results of this study. A second study (Daniel et
    al., 1992a), however, showed that chloral hydrate administered in
    drinking-water to a group of 40 male B6C3F1 mice for 104 weeks at
    1 g/litre (166 mg/kg of body weight per day) resulted in a 71%
    incidence of hepatic tumours (combined adenomas and carcinomas). The
    fact that much higher doses were required to induce tumours in the
    same hybrid mouse in this study raises further questions about the
    Rijhsinghani et al. (1986) study; however, mice of this age are known
    to be very sensitive to tumour initiators (Vesselinovich et al.,
    1974). However, the Daniel et al. (1992a) study does clearly indicate
    that chloral hydrate is capable of inducing tumours in B6C3F1 mice
    when the mice are subjected to a lifetime exposure.a


              

    a  After the Task Group meeting, three new studies in B6C3F1 mice
    on the carcinogenicity of chloral hydrate appeared (NTP, 2000a,b;
    George et al., in press). In the first (NTP, 2000a), no carcinogenic
    effect was observed following the administration of a single dose of
    chloral hydrate (the dose was up to 5 times higher than that used in
    the study of Rijhsinghai et al., 1986), while in the two other
    studies, males had an increased incidence of hepatic tumours after
    life-time exposure (NTP, 2000b; George et al., in press). In the NTP
    life-time study (NTP, 2000b), a slightly elevated incidence of
    pituitary adenomas, of borderline statistical significance, was
    observed in female mice.

         The question of whether chloral hydrate itself contributes to its
    carcinogenic effects is critical because at least two of its
    metabolites, TCA and DCA, are comparatively potent inducers of hepatic
    tumours in B6C3F1 mice. This question can be resolved only by
    demonstrating (i) that the clastogenic effects of chloral hydrate play
    a role in the development of tumours or (ii) that TCA, DCA or a
    combination of both chemicals are produced in sufficient quantities to
    completely account for the induction of liver tumours without
    significant contribution from earlier, more reactive metabolites. This
    question is more critically assessed in the next section.

    4.3.1.4  Comparative metabolism and pharmacokinetics

         The metabolism of chloral hydrate has received considerable
    attention over the years because of its extensive use as a
    sedative-hypnotic (Marshall & Owens, 1954; Kaplan et al., 1967;
    Gessner & Cabana, 1970; Cabana & Gessner, 1970; Sellers et al., 1972;
    Garrett & Lambert, 1973; Mayers et al., 1991). A series of older
    studies provide a still valid general picture of the conversion of
    chloral hydrate to its two major metabolites, trichloroethanol and
    TCA. More recent studies have focused more specifically on species
    differences in this metabolism and have begun to focus on minor
    metabolic pathways.

         Figure 6 provides a simplified scheme of chloral hydrate
    metabolism. The reader is referred to the appropriate sections of this
    document to evaluate the further metabolism of the products TCA and
    DCA.

         The major fate of chloral hydrate is to undergo reduction to
    trichloroethanol, with a smaller, but significant, fraction being
    oxidized to TCA. Initially, formation of trichloroethanol is favoured
    because the redox potential within cells  in vivo favours reduction
    (Kawamoto et al., 1987). This initial tendency is accentuated by a
    rapid glucuronidation of the trichloroethanol that is formed (Marshall
    & Owens, 1954). As is pointed out below, this may be a key feature in
    the interspecies differences in chloral hydrate metabolism. With time,

                (continued)

    References:
    NTP (2000a) Toxicology and carcinogenesis studies of chloral hydrate
    in B6C3F1 mice (gavage studies).  Research Triangle Park, North
    Carolina, US department of Health and Human Services, National
    Toxicology Program (NTP-TR-502).

    NTP (2000b) Toxicology and carcinogenesis studies of chloral hydrate
    (ad libitum and dietary controlled) in male B6C3F1 mice (gavage
    study).  Research Triangle Park, North Carolina, US department of
    Health and Human Services, National Toxicology Program (NTP-TR-503).

    George MH, Kilburn S, Moore T & DeAngelo AB (in press)  The
    carcinogenicity of chloral hydrate administered in the drinking water
    to the male B6C3F1 mouse and F344/N rat. Toxicol Pathol.

    FIGURE 7

    however, more TCA is formed as a result of enterohepatic circulation
    of the trichloroethanol glucuronide (Stenner et al., 1996). The
    glucuronide is hydrolysed to trichloroethanol, which can be oxidized
    to TCA, with chloral hydrate as an intermediate. Under physiological
    conditions, the formation of TCA is for all practical purposes
    irreversible, and the net conversion of chloral hydrate to TCA will
    continue as long as there are significant amounts of trichloroethanol
    entrained in the enterohepatic circulation.

         The production of these major products of chloral hydrate
    metabolism is relatively well understood. However, the exact source of
    a number of minor metabolites is less well understood. The reaction
    rates and mechanisms involved are just beginning to be studied.
    Understanding these mechanisms will be key to understanding whether
    the effects of chloral hydrate can be attributed primarily to its
    conversion to chemicals of established toxicological properties, such
    as trichloroethanol, DCA and TCA, or whether the activities of
    reactive intermediates must also be considered. Formation of DCA
    probably requires radical formation, but it is not clear whether the
    radical would be formed from trichloroethanol, chloral hydrate or TCA.
    If DCA is derived from the first two compounds, the
    dichloroacetaldehyde that would result as an intermediate could pose
    some toxicological problems as well. Moreover, Ni et al. (1996)
    suggested from their ESR data that a trichloromethyl radical is formed
    from chloral hydrate. Such an intermediate could contribute to the
    toxicological effects of chloral hydrate as well.

         Available pharmacokinetic data have focused upon the relative
    role of trichloroethanol as the metabolite of chloral hydrate that is
    responsible for its central nervous system-depressant activity
    (Butler, 1948, 1949; Marshall & Owens, 1954; Garrett & Lambert, 1973).
    Conversely, a variety of metabolites have been suggested as
    responsible for the toxic effects of chloral hydrate and its ability
    to induce liver cancer in mice, in particular (Daniel et al., 1992a).
    The critical question is whether an early reactive metabolite -- e.g.,
    the free radicals identified by Ni et al. (1996) -- induces
    clastogenic effects that are important to tumour development. The
    competing, but not necessarily exclusive, hypothesis is that the
    clearly hepatocarcinogenic metabolites TCA and DCA are produced in
    sufficient quantity to account for some or all of the liver cancer
    that results from chloral hydrate treatment.

         While TCA and trichloroethanol are well established metabolites
    of chloral hydrate, DCA has been only recently recognized as a
    potentially important metabolite with respect to liver tumour
    induction. Part of the difficulty is that DCA is much more rapidly
    metabolized than either TCA or trichloroethanol, and peak
    concentrations would be expected to be significantly lower. This first
    became apparent when DCA appeared to be produced in significant
    quantities from trichloroethylene administered to B6C3F1 mice
    (Templin et al., 1993). Chloral hydrate is the first stable metabolite
    of trichloroethylene metabolism (Cole et al., 1975); thus, these data
    prompt an examination of the role that chloral hydrate plays in the
    formation of DCA. In contrast with these results in mice, DCA was not

    detectable in rats and dogs administered similar doses (Templin et
    al., 1995). It became apparent that some of the DCA that was measured
    in these studies may have arisen artefactually through dehalogenation
    of TCA under acid conditions in fresh blood (Ketcha et al., 1996). As
    a consequence of this series of findings, newer studies have examined
    whether DCA is produced from chloral hydrate in a number of species.

         Recent results of Abbas et al. (1996) showed that doses of 10 and
    100 mg of chloral hydrate per kg of body weight result in
    approximately 2.4 and 10 g of DCA per ml of blood. However, it must
    be noted that a variety of artefacts of DCA formation from TCA in
    blood suggest that these results need to be viewed with caution
    (Ketcha et al., 1996). Indeed, Merdink et al. (1998) found that DCA
    was not measurable in mice dosed with 50 mg of chloral hydrate per kg
    of body weight.

    4.3.1.5  Mode of action

         The mechanism of action involved in chloral hydrate-induced liver
    tumours in mice remains to be established. Clearly, chloral hydrate is
    converted to at least one metabolite, TCA, that appears to act as a
    peroxisome proliferator. There is some possibility that it is
    converted to DCA, a compound that acts primarily as a tumour promoter
    (Stauber & Bull, 1997). On the other hand, chloral hydrate is distinct
    from these other two compounds in that it appears to be clastogenic
     in vivo, but at very high doses. The question is, does this
    clastogenic activity play a role in tumorigenesis? The fact that
    chloral hydrate appears to produce only hepatic tumours in mice
    parallels the species specificity of TCA and suggests that other
    activities are perhaps not involved.

    4.3.2  Halogenated aldehydes and ketones other than chloral hydrate

    4.3.2.1  General toxicological properties and information on 
             dose-response in animals

    1)   Haloaldehydes

         Relatively few data are available to describe acute, short-term
    or chronic toxicities for the haloacetaldehydes. In general, aldehydes
    are irritant chemicals, and substitution of chlorine generally
    increases this irritancy. However, it is unlikely that irritant
    effects will occur at concentrations that are encountered in
    drinking-water. Chloroacetaldehyde is an example of a haloaldehyde for
    which some data exist, although it has not been commonly found in
    drinking-water. Concentrations of chloroacetaldehyde as low as 0.02%
    produced intradermal irritation, 7.5% produced rather severe dermal
    irritation and 0.03% irritated the eyes of rabbits (Lawrence et al.,
    1972).

         When administered systemically, halogenated aldehydes are quite
    toxic. Again because of a lack of appropriate data, chloroacetaldehyde
    will be used for illustrative purposes. Lawrence et al. (1972)
    reported the oral LD50 in Sprague-Dawley rats to be 89 and 103 mg/kg

    of body weight in males and females, respectively. In male ICR mice,
    the oral LD50 was found to be 82 mg/kg of body weight. In longer-term
    investigations, Lawrence and co-workers (1972) utilized
    intraperitoneal injections or inhalation as the method of
    administration. Intraperitoneal injections of 2.2 or 4.4 mg/kg of body
    weight per day for 30 days in male Sprague-Dawley rats produced
    significant decreases in haemoglobin, haematocrit and erythrocytes at
    the highest dose. This was consistent with its ability to induce
    haemolysis of erythrocytes at concentrations of 0.2 mol of
    chloroacetaldehyde per litre and above. However, the validity of this
     in vitro observation as a dependable indicator of  in vivo effects
    is suspect because of the extremely high concentrations that were
    utilized. The 4.4 mg/kg of body weight dose produced reductions in
    body weight and induced significant increases in the organ to body
    weight ratios for the brain, gonads, heart, kidneys, lungs and spleen.
    These effects appeared to be largely the result of reduced body weight
    rather than changes in absolute organ weights.

         Lawrence et al. (1972) pointed out the importance of route of
    administration to the toxicity of chloroacetaldehyde. If administered
    intraperitoneally, it is 10-30 times as potent as its metabolite
    2-chloroethanol. However, the toxicity of the two compounds is more or
    less equivalent when they are administered orally, and 2-chloroethanol
    is about 4 times as toxic as chloroacetaldehyde when applied
    topically. These differences are most likely attributed to relative
    rates of absorption versus metabolic conversion and the many
    non-specific reactions in which chloroacetaldehyde would be expected
    to be involved in the gastrointestinal tract and on the skin. This
    issue should be reflected in the systemic toxicities of halogenated
    aldehydes, in general.

         A 104-week study of chloroacetaldehyde utilizing drinking-water
    as the mode of administration was conducted in male B6C3F1 mice
    (Daniel et al., 1992a). Only a single concentration was used
    (100 mg/litre), which yielded an average dose of 17 mg/kg of body
    weight (i.e., larger than administered to rats intraperitoneally).
    This dose did not lead to excessive mortality or depress body weight
    gains. It did not affect weights of the liver, kidneys, testes or
    spleen. There was no remarkable non-tumour pathology in 40 tissues
    that were sampled and examined microscopically in five of the animals
    that were serially sacrificed during the course of the experiment. An
    apparent increase in the incidence of liver histopathological change
    was described as hepatocellular necrosis, hyperplasia and cytomegaly.
    However, these effects were very mild and of doubtful significance as
    compared with the same types of pathology in control mice.

         Sood & O'Brien (1993) examined the effects of chloroacetaldehyde
    in isolated rat hepatocytes. A concentration of 0.5 mmol/litre was
    found to be cytotoxic, whereas 0.2 mmol/litre was without apparent
    effect. The cytotoxicity could be essentially abolished by
    dithiothreitol in the incubation media. The requirement for high
    concentrations in this  in vitro experiment and the apparent lack of
    effect (albeit administered to a different species) at relatively high
    concentrations in the animals' drinking-water (100 mg/litre) suggest

    that there is little concern about hepatotoxicity at the very low
    concentrations that might be expected to be found in chlorinated
    drinking-water.

    2)   Haloketones

         Toxicological data in experimental animals for the haloketones
    are extremely limited. The halopropanones are the most commonly
    studied group of this class, but most of the work has been directed at
    mutagenic effects of the chemicals.

         Laurie et al. (1986) studied the effects of 1,1- and 1,3-DCPN in
    CD-1 mice. 1,1-DCPN was administered in paraffin, and 1,3-DCPN was
    administered as an aqueous solution. 1,1-DCPN significantly increased
    the levels of ASAT, ALAT and LDH at doses of 325 mg/kg of body weight.
    1,3-DCPN was evaluated to a maximum dose of 20 mg/kg of body weight
    and appeared to be without effect on the serum enzymes. However, the
    LD50 for 1,3-DCPN was stated to be 25 mg/kg of body weight,
    indicating that the liver was probably not the critical target organ
    for this compound, at least with acute treatment. At doses of 130
    mg/kg of body weight and greater, 1,1-DCPN significantly depressed
    hepatic GSH levels. Again, 1,3-DCPN was without effect at the dose of
    20 mg/kg of body weight. Most of the decrease in GSH produced by
    1,1-DCPN was observed in the post-mitochondrial cellular fraction as
    opposed to the mitochondrial fraction.

         A major thrust of the Laurie et al. (1986) paper was to determine
    the extent to which 1,1- and 1,3-DCPN modified the toxicity of carbon
    tetrachloride. Carbon tetrachloride was administered at doses ranging
    from 0.02 to 1.0 ml/kg of body weight. The 0.02 ml/kg of body weight
    dose produced an elevation of serum enzymes on its own. However, when
    carbon tetrachloride was administered 4 h after 1,1-DCPN, the dose of
    1,1-DCPN that was required to significantly increase serum enzyme
    levels was decreased from 325 to 130 mg/kg of body weight. Since a
    doubling of the dose of 1,1-DCPN would have produced a similar
    response and the dose of carbon tetrachloride was above a threshold
    response level, the interaction between carbon tetrachloride and
    1,1-DCPN would seem to be no more than additive. Inhibition of the
    toxicological effects of carbon tetrachloride in a dose-related manner
    was observed when 1,3-DCPN was administered prior to carbon
    tetrachloride.

         Merrick et al. (1987) studied the cytotoxic effects of
    chloropropanone (CPN), 1,1-DCPN and 1,3-DCPN on isolated hepatocytes
    of male Sprague-Dawley rats. The chloropropanones were all shown to
    react with GSH in solution. At concentrations of 10 mmol/litre, the
    rate of GSH reaction was most rapid with 1,3-DCPN, followed by CPN and
    then 1,1-DCPN. This reactivity paralleled the ability of the compounds
    to induce cytotoxicity in isolated hepatocytes. Significant increases
    in ASAT release were observed with 1,3-DCPN at 0.5 mmol/litre, CPN at
    1 mmol/litre and 1,1-DCPN at 5 mmol/litre. As would be expected, GSH
    depletion was observed at concentrations of 1,3-DCPN as low as 0.1
    mmol/litre. The other two compounds were significantly less active in

    depleting GSH, but were of approximately equivalent potency with one
    another. 

         A study of 1,1,1-TCPN was conducted in Sprague-Dawley rats by
    Daniel et al. (1993b). Acute, 10-day and 90-day experiments were
    performed. 1,1,1-TCPN was administered in corn oil by gavage. Doses in
    the 10-day study were 0, 16, 48, 161 or 483 mg/kg of body weight per
    day. In the 90-day study, doses of 30, 90 or 270 mg/kg of body weight
    per day were administered. In the 10-day study, 8 out of 10 male and 7
    out of 10 female rats died at 483 mg/kg of body weight per day before
    the conclusion of the treatments. Two male rats also died at 161 mg/kg
    of body weight per day. Although there was not a significant effect on
    body weight in the survivors, there was a 10% increase in liver to
    body weight ratios at 161 mg/kg of body weight per day in both male
    and female rats. Evidence of hyperkeratosis was found in the
    forestomach of both male and female rats treated at doses of 48 mg/kg
    of body weight per day and greater. No adverse effect was observed at
    16 mg/kg of body weight per day.

         Increases in relative liver weight were observed in the 90-day
    experiment at 270 mg/kg of body weight per day in male rats, but not
    at 90 mg/kg of body weight per day. Ataxia was reported in both sexes
    at the 270 mg/kg of body weight per day dose level. Increases in the
    incidence of forestomach lesions were observed in both sexes at 90 and
    270 mg/kg of body weight per day, with the most frequent observation
    being hyperkeratosis followed by acanthosis. The overall NOAEL in the
    10-day and 90-day studies is 30 mg/kg of body weight per day.

         In summary, the toxicological effects of the halopropanones
    provide evidence that some of the representatives of this class are
    highly toxic, with acute lethal doses being as low as 25 mg/kg of body
    weight. The gastrointestinal tract and liver appear to be key target
    organs for some members of the class. However, no target organ was
    identified for the most acutely toxic of the group, 1,3-DCPN.

    4.3.2.2  Toxicity in humans

         No data on the effects of either halogenated aldehydes or
    halogenated ketones on human subjects were identified.

    4.3.2.3  Carcinogenicity and mutagenicity

         There are considerable data on the mutagenic properties of
    various halogenated aldehydes and ketones. A comparison of the
    mutagenic activity, measured in  Salmonella typhimurium tester
    strains, of the halogenated aldehydes found in drinking-water with
    that of chloroacetaldehyde is summarized (not comprehensively) in
    Table 19. Table 20 provides a similar summary of results with
    halogenated ketones. These data are presented to demonstrate that the
    activity of these compounds is not dissimilar from that observed with
    chloroacetaldehyde, a metabolite of a large number of carcinogenic
    chemicals, including vinyl chloride. Therefore, it will be used as a
    prototype for the class. This is not to suggest that this is an
    adequate substitute for appropriate data for the other compounds. It

    must be recognized that the data available for this class are
    completely inadequate for making substantive estimates of the impact
    of these chemicals on human health; in particular, mutagenicity data
    in bacterial systems do not necessarily reflect activity  in vivo. 

    1)   Haloaldehydes

         Chloroacetaldehyde has been extensively studied with respect to
    the types of interactions that it has with DNA. The adducts formed in
    animals treated with vinyl chloride are the same as those produced
    with chloroacetaldehyde (Green & Hathway, 1978). The cyclic etheno
    adducts formed with cytosine and adenine seem particularly important
    in mutagenic responses observed with chloroacetaldehyde (Spengler &
    Singer, 1988; Jacobsen et al., 1989). 

         There are a limited number of studies that have examined the
    carcinogenic properties of chloroacetaldehyde in rather specialized
    test systems. Van Duuren et al. (1979) examined the carcinogenic
    activity of chloroacetaldehyde in mouse skin initiation/promotion
    studies, with subcutaneous injection and by stomach tube. In the
    initiation/promotion assay, chloroacetaldehyde was applied to the skin
    of 30 female Ha:ICR Swiss mice, at a dose of 1.0 mg per application
    per mouse, 3 times weekly for up to 581 days, or as a single dose
    followed by 2.5 g of TPA 3 times weekly for 576 days. No evidence of
    increased skin tumour yield was found. The intragastric treatments
    involved administration of 0.25 mg per mouse per week (1.8 mg/kg of
    body weight per day if a 20-g body weight is assumed). In this case,
    sections of lung, liver and stomach were taken for histopathological
    examination. No signs of increased tumour incidence were found. A
    fourth experiment involved the subcutaneous injection in 30 mice of
    0.25 mg per mouse (1.8 mg/kg of body weight per day, assuming a 20-g
    body weight for the mouse). Microscopic examination of sections of the
    liver and injection sites revealed no evidence of increased tumour
    yield.

         A second group of mouse skin initiation/promotion experiments
    with chloroacetaldehyde were conducted by Zajdela et al. (1980).
    Single doses of 0.05, 0.1, 1.0 or 2.5 mg of chloroacetaldehyde
    dissolved in acetone were applied to male and female XVIInc/Z mice
    (20-28 per group). This was followed by application of TPA at 2 g, 3
    times weekly for 42 weeks. There was no significant difference between
    mice receiving TPA alone and those that had been initiated by
    chloroacetaldehyde.

         There appears to be only one study that examined the carcinogenic
    activity of halogenated aldehydes administered in drinking-water over
    a lifetime (Daniel et al., 1992a). Male B6C3F1 mice treated with 0.1
    g of chloroacetaldehyde per litre of drinking-water for 104 weeks were
    found to have an incidence of eight hepatocellular carcinomas in
    26 mice examined (31%). In addition, 8% of these mice were diagnosed
    as having hepatocellular adenoma, and another 8% were found to have
    hyperplastic nodules. This compared with two carcinomas in 20 control
    mice examined (10%), one with adenoma (5%) and no hyperplastic nodules
    (0%). In a comparison with previous studies, the experiment utilized a

        Table 19. Mutagenic activity of halogenated aldehydes produced by chlorination in 
              the Salmonella/microsome assay

                                                                                         
    Compound                      Strain    Net revertants/plate   Reference
                                                                
                                            -S9       +S9
                                                                                         
    Chloroacetaldehyde            TA100     440       ~330         Bignami et al. (1980)
    2-Chloropropenal              TA100     135       49           Segall et al. (1985)
    2-Bromopropenal               TA100     1140      108          Segall et al. (1985)
    2-Bromopropenal               TA100     400                    Gordon et al. (1985)
    2,3-Dibromopropanal           TA100     300                    Gordon et al. (1985)
    2,3-Dichloropropenal          TA100     91        5            Segall et al. (1985)
    3,3-Dichloropropenal          TA100     0.7                    Meier et al. (1985a,b)
    2,3,3-Trichloropropenal       TA100     224                    Rosen et al. (1980)
    3-Chloro-2-butenal            TA100     68        42           Segall et al. (1985)
    3-Bromo-2-butenal             TA100     108       39           Segall et al. (1985)
    2-Bromo-3-methyl-2-butenal    TA100     <0.5      <0.5         Segall et al. (1985)
                                                                                         


    Table 20. Mutagenic activity of halogenated ketones produced by chlorination in the 
              Salmonella/microsome assay

                                                                                           
    Compound                       Strain     Net revertants/plate   Reference
                                                                   
                                              -S9      +S9
                                                                                           

    CPN                            TA100      NMa      NM            Merrick et al. (1987)
    1,1-DCPN                       TA100      0.04                   Meier et al. (1985a)
    1,3-DCPN                       TA100      25.2                   Meier et al. (1985a)
    1,1,1-TCPN                     TA100      0.12                   Meier et al. (1985a)
    1,1,3-TCPN                     TA100      3.9                    Douglas et al. (1985)
    1,1,3,3-Tetrachloropropanone   TA100      1.5                    Meier et al. (1985a)
    Pentachloropropanone           TA100      0.86                   Meier et al. (1985a)
                                                                                           

    a NM = non-mutagenic.
    

    significantly higher dose (17 mg/kg of body weight per day) as well as
    a continuous treatment. It is possible that these tumours were induced
    by the genotoxic properties of the chemical.

         Robinson et al. (1989) examined the ability of four other
    halogenated aldehydes to act as tumour initiators in the skin of
    Sencar mice: 2-chloropropenal, 2-bromopropenal, 3,3-dichloropropenal
    and 2,3,3-trichloropropenal. The compounds were administered topically
    in six divided doses over a 2-week period (total doses were

    600-2400 mg/kg of body weight). Two weeks after the final initiating
    dose, TPA was applied at a dose of 1 g, 3 times weekly for 20 weeks.
    Both 2-chloropropenal and 2-bromopropenal significantly increased
    tumour yield at 24 weeks and significantly increased the yield of
    squamous cell carcinomas at 52 weeks at total topical doses of
    1200 mg/kg of body weight and above. Both the benign tumour and
    malignant tumour yields were greater with 2-bromopropenal than with
    2-chloropropenal. An experiment utilizing oral administration of these
    compounds during the initiation period was included in this study.
    Oral administration of 2-chloropropenal did not produce consistent,
    dose-related responses. However, there appeared to be a substantial
    increase in skin tumour yield at an oral dose of 300 mg of
    2-bromopropenal per kg of body weight (19 tumours in 38 mice [50%] vs.
    20 tumours in 110 control mice [18%]).

         On the basis of these studies, it must be concluded that there is
    a potential carcinogenic hazard associated with the halogenated
    aldehydes. Only a single compound, chloroacetaldehyde, was evaluated
    as a carcinogen in a lifetime study, and only one dose level was
    studied. It appears to be more potent as a carcinogen than the
    corresponding THM and HAA by-products. Many members of the class are
    mutagenic, and chloroacetaldehyde, at least, appears to produce
    tumours in the liver at less than cytotoxic doses. Based upon the
    comparison between 2-chloropropenal and 2-bromopropenal, there is some
    reason to believe that the brominated by-products are more potent than
    the corresponding chlorinated by-products. Therefore, concern must be
    expressed over disinfection processes that activate bromide, as well
    as those that simply chlorinate. However, the currently available data
    are not sufficient to allow the hazards associated with these
    compounds to be estimated.

    2)   Haloketones

         A number of halopropanones have been tested in mutagenesis
    assays. To facilitate comparison of their relative potencies, selected
    results from assays that were conducted in  Salmonella typhimurium 
    tester strain TA100 were incorporated into Table 20. For the most
    part, the data selected for this table were abstracted from papers in
    which more than one haloketone was evaluated, rather than being
    selected because they were identified as the best value for each
    individual chemical that exists in the literature. Some of the
    compounds have been shown to be active in other  Salmonella tester
    strains and other mutagenesis and clastogenesis assays. There was
    little to be gained from an exhaustive review of this literature, so
    further consideration of the mutagenic activity will be limited to
    those systems that extended evaluations to other end-points  in 
     vitro or attempted to confirm  in vitro observations  in vivo. 

         1,3-DCPN was found to induce SCEs in V79 cells at concentrations
    as low as 0.002 mmol/litre (von der Hude et al., 1987). Blazak et al.
    (1988) found that 1,1,1-TCPN and 1,1,3-TCPN were able to act as
    clastogens in CHO cells  in vitro. Structural aberrations were
    produced at a 1,1,3-TCPN concentration of 1.5 g/ml, whereas a
    concentration of 23 g/ml was required for a similar response to

    1,1,1-TCPN. However, the dose-response for 1,1,3-TCPN was limited by
    cytotoxicity. Experiments were also conducted focusing on the ability
    of 1,1,1-TCPN and 1,1,3-TCPN to induce micronuclei in polychromatic
    erythrocytes and to induce sperm head abnormalities in mice  in 
     vivo. 1,1,1-TCPN was found to be negative in both assays in the dose
    range 75-300 mg/kg of body weight, whereas 1,1,3-TCPN was negative in
    the range 3-12 mg/kg of body weight.

         Robinson et al. (1989) tested CPN, 1,1-DCPN, 1,3-DCPN, 1,1,1-TCPN
    and 1,1,3-TCPN as initiators in the skin of Sencar mice. The compounds
    were administered by topical application in acetone with total doses
    that ranged from 37.5 to 4800 mg/kg of body weight (doses of 600 mg/kg
    of body weight and above were administered in six equal doses over a
    2-week period to avoid cytotoxic or lethal effects of the compounds).
    Other groups of animals treated with the chemicals were also
    administered similar doses by intragastric intubation. The initiating
    treatments were followed by a promotion schedule that involved the
    topical application of 1 g of TPA 3 times weekly for 20 weeks. Tumour
    counts were reported at 24 weeks; if the incidence was elevated within
    this time period, the mice were held until 52 weeks on study prior to
    sacrifice, and histological evaluations of the tumours were made.
    Among the haloketones, only 1,3-DCPN was found to produce a
    dose-related increase in tumour incidence. A single topical dose of
    37.5 mg/kg of body weight was sufficient to initiate skin tumours, and
    the response increased progressively as doses were increased to 150
    mg/kg of body weight. At higher doses, the response decreased in
    magnitude. Splitting the 300 mg/kg of body weight dose into six equal
    doses over a 2-week period increased the tumorigenic response relative
    to a single dose of 300 mg/kg of body weight. However, this response
    was also attenuated, as the multiple-dose schedule utilized higher
    doses. This attenuation of the response was particularly marked in
    total tumour yields, which included many benign tumours. It was less
    effective in limiting the yield of squamous cell carcinomas.

         In conclusion, the carcinogenic activity of the DBPs in the
    haloaldehyde and haloketone classes, with the exception of
    chloroacetaldehyde, has not been evaluated in lifetime studies in
    experimental animals. However, other tests confirm that they have
    carcinogenic properties. 1,3-DCPN was the most potent tumour initiator
    in both classes of DBPs. A single dose of 75 mg/kg of body weight
    produces a total tumour yield equivalent to that produced by 1200 mg
    of 2-bromopropenal, the most potent of the haloaldehydes, per kg of
    body weight. 2-Bromopropenal is about 40 times as potent as 1,3-DCPN
    as a mutagen. The other halopropanones do not appear to be capable of
    acting as tumour initiators in the mouse skin. 

    4.3.2.4  Comparative pharmacokinetics and metabolism

         No information was identified in the available literature.

    4.3.2.5  Mode of action

         The data available indicate that these two groups of chemicals
    contain compounds that possess mutagenic activity. As these effects
    have been identified in  in vitro or bacterial test systems, there is
    no assurance that this is the manner in which they contribute to
    toxicity or carcinogenicity. A few chemicals have been shown to be
    initiators in the mouse skin, but it is not clear whether that would
    be a target organ as a result of chronic ingestion of these chemicals.
    Other chemicals appear to have activities that could contribute less
    directly to the induction of cancer, particularly as cytotoxic
    compounds. It is clear from the limited data available that it would
    be inappropriate to try to generalize data from only a few examples to
    these two larger classes of DBPs.

    4.4  Haloacetonitriles

    4.4.1  General toxicological properties and information on 
           dose-response in animals and humans

         The HANs are discussed in a single section of this document
    because the toxicological data on them are quite limited. The
    dihaloacetonitriles (DHAN) -- DCAN, BCAN and DBAN -- are the most
    important in terms of concentrations found in chlorinated
    drinking-water. However, there are limited data on bromoacetonitrile
    (BAN), chloroacetonitrile (CAN) and trichloroacetonitrile (TCAN) that
    are included for completeness.

         Hayes et al. (1986) examined the general toxicological effects of
    DCAN and DBAN in male and female ICR mice and CD rats. In mice, the
    acute oral (by gavage in corn oil) LD50 was reported to be 270
    (males) and 279 (females) mg/kg of body weight for DCAN and 289
    (males) and 303 (females) mg/kg of body weight for DBAN. In rats, the
    LD50 was found to be 339 (males) and 330 (females) mg/kg of body
    weight for DCAN and 245 (males) and 361 (females) mg/kg of body weight
    for DBAN. Hussein & Ahmed (1987) found somewhat lower oral LD50s in
    rats: BAN, 25.8 mg/kg of body weight; DBAN, 98.9 mg/kg of body weight;
    CAN, 152.8 mg/kg of body weight; and DCAN, 202.4 mg/kg of body weight.
    These latter data were reported only in abstract form, and the vehicle
    used was not indicated. As discussed below, some of the toxicological
    responses to chemicals in this class appear to depend on the nature of
    the vehicle in which they were administered.

         DCAN and DBAN were also studied over 14- and 90-day treatment
    intervals (Hayes et al., 1986). DCAN dissolved in corn oil was
    administered to male and female CD rats by gavage at 12, 23, 45 or 90
    mg/kg of body weight per day for 14 days and at 8, 33 or 65 mg/kg of
    body weight per day for 90 days. DBAN was administered to male and
    female CD rats at daily doses of 23, 45, 90 or 180 mg/kg of body
    weight per day for 14 days and at 6, 23 or 45 mg/kg of body weight per
    day for 90 days. Increased mortality was produced at 33 mg of DCAN per
    kg of body weight per day and at 45 mg of DBAN per kg of body weight
    per day in the 90-day studies. Body weight was decreased and lower
    weights and organ to body weight ratios were observed for spleen and

    gonads with doses of 65 mg of DCAN per kg of body weight per day and
    above. The NOAELs for DCAN were 45 mg/kg of body weight per day for 14
    days and 8 mg/kg of body weight per day for 90 days of exposure. The
    NOAELs for DBAN were 23 mg/kg of body weight per day at 90 days and 45
    mg/kg of body weight per day at 14 days. No serum chemistry changes
    indicative of adverse effects were seen with either compound at
    sublethal doses.

    4.4.2  Reproductive and developmental toxicity

         Smith et al. (1987) examined the effect of CAN, DCAN, TCAN, BCAN
    and DBAN on female reproduction in an  in vivo teratology screening
    test in Long-Evans hooded rats. DCAN and TCAN at doses of 55 mg/kg of
    body weight administered in tricaprylin by gavage from day 7 to day 21
    of gestation significantly reduced the percentage of females
    delivering viable litters, increased resorption rates and reduced
    maternal weight gain. BCAN and DBAN at the same dose were without
    effect. All of the HANs reduced the mean birth weight of pups, and the
    DHANs reduced the postnatal weight gain till the fourth day after
    birth. Postnatal survival was reduced with DCAN and TCAN but not with
    BCAN or DBAN. These pups continued to display reduced body weights
    into puberty. BCAN also resulted in significantly depressed weights at
    puberty, although the effect was smaller than that observed with DCAN
    or TCAN.

         The hydra assay system for developmental toxicity has also been
    used to screen some of the HANs (Fu et al., 1990). Both DBAN and TCAN
    were found to be of the same general order of toxicity to adult and
    embryonic animals. Based on these findings, the authors predicted that
    DBAN and TCAN would not be teratogenic at non-maternally toxic doses.

         The developmental toxicity of DCAN was followed up in full-scale
    teratology studies (Smith et al., 1989b). In this case, DCAN dissolved
    in tricaprylin was administered to Long-Evans rats at doses of 0, 5,
    15, 25 or 45 mg/kg of body weight per day from day 6 to day 18 of
    gestation. Embryolethality and fetal resorptions were statistically
    significant at 25 and 45 mg/kg of body weight per day. The highest
    dose was also maternally toxic. Soft tissue anomalies, including an
    intraventricular septal defect in the heart, hydronephrosis, fused
    ureters and cryptorchidism, were observed at this dose. Skeletal
    abnormalities (fused and cervical ribs) were produced in a
    dose-dependent manner and were significantly increased at 45 mg/kg of
    body weight per day. A NOAEL was found to be 15 mg/kg of body weight
    per day.

         TCAN was evaluated in two teratology studies by the same
    laboratory (Smith et al., 1988; Christ et al., 1996). The first of
    these studies utilized tricaprylin as the vehicle, whereas the second
    utilized corn oil. In the first study, embryolethality was observed at
    doses as low as 7.5 mg/kg of body weight per day. Doses of 15 mg/kg of
    body weight per day and above produced soft tissue abnormalities,
    including fetal cardiovascular anomalies (Smith et al., 1988). TCAN
    administered in a corn oil vehicle produced cardiovascular defects at
    55 mg/kg of body weight per day (Christ et al., 1996). The effects

    observed at this dose were found at significantly lower incidence than
    observed in the 15 mg/kg of body weight per day dose of the previous
    study. The abnormalities were milder, being simply positional
    (laevocardia), instead of the interventricular septal defect and a
    defect between the ascending aorta and right ventricle that were
    observed with the tricaprylin vehicle. On the other hand, more
    skeletal defects were observed with the corn oil vehicle. The authors
    attributed these differences to an interaction between the tricaprylin
    vehicle and TCAN. However, tricaprylin and corn oil are not
    representative of drinking-water exposure. It is not possible to
    determine whether the results obtained with tricaprylin or the results
    obtained with corn oil provide the most valid test. The results could
    be just as readily ascribed to an inhibition of the effects of the
    corn oil vehicle. As a consequence, these data present somewhat of a
    difficulty in interpreting results for all of the HANs that have been
    tested, because the only vehicle in which most have been evaluated was
    tricaprylin.

    4.4.3  Carcinogenicity and mutagenicity

         IARC has evaluated BCAN, CAN, DBAN, DCAN and TCAN and concluded
    that there is inadequate evidence for their carcinogenicity in
    experimental animals. No data were available on their carcinogenicity
    in humans. Consequently, these HANs were assigned to Group 3: the
    agent is not classifiable as to its carcinogenicity to humans (IARC,
    1991, 1999).

         Bull et al. (1985) tested the ability of CAN, DCAN, TCAN, BCAN
    and DBAN to induce point mutations in the  Salmonella/microsome
    assay, to induce SCEs in CHO cells  in vitro, to produce micronuclei
    in polychromatic erythrocytes in CD-1 mice and to act as tumour
    initiators in the skin of Sencar mice.

         DCAN produced a clear increase in mutagenic activity in
     Salmonella typhimurium strains TA1535 and TA100. This response was
    not altered by the inclusion of the S9 system to metabolically
    activate the compound, if needed. BCAN also produced a positive
    response at low doses, but the dose-response curve was interrupted at
    high doses by cytotoxicity. In this case, the inclusion of the S9
    fraction appeared to simply allow the bacteria to survive cytotoxic
    effects. This perhaps arose through a non-specific inactivation of the
    electrophilic character of the compound. BCAN produced a similar, but
    less marked, trend in strain TA100. The other HANs were negative in
    the  Salmonella/microsome assay.

         All of the HANs tested increased the frequency of SCEs in CHO
    cells. The potencies in the absence of S9 were DBAN > BCAN > DCAN
    approx. or equiv. TCAN > CAN. As in the  Salmonella/microsome assay,
    the addition of S9 allowed higher doses to be tested rather than
    modifying the response to a given concentration. In contrast, none of
    the HANs was found to induce micronuclei in CD-1 mice  in vivo. 

         The HANs were tested for their ability to initiate tumours in the
    skin of Sencar mice (Bull et al., 1985). In this experiment, the HANs
    were administered topically to the shaved backs of the mice in six
    doses over a 2-week period. Two weeks following the last initiation
    treatment, TPA dissolved in acetone was applied topically at a dose of
    1 g per mouse, 3 times weekly for 20 weeks. The total initiating
    doses were 1200, 2400 and 4800 mg/kg of body weight. Significant
    increases in skin tumours were observed with CAN, TCAN, BCAN and DBAN.
    DBAN produced the greatest response at a dose of 2400 mg/kg of body
    weight, but the response decreased in magnitude as the dose was
    increased to 4800 mg/kg of body weight. The shape of this
    dose-response curve was checked in a repetition of the experiment, and
    the results were virtually identical. It was postulated that the
    attenuated response at the higher dose was caused by the cytotoxicity
    to initiated cells. BCAN also increased the incidence of skin tumours,
    but no significant increase in tumour incidence was induced by DCAN.
    The carcinogenicity of the DHANs in mouse skin was seen to
    progressively increase as bromine was substituted for chlorine in the
    compound. On the other hand, CAN appeared to be among the more potent
    of the HANs in initiating skin tumours. TCAN gave inconsistent
    results.

         CAN, TCAN and BCAN produced small, but significant, increases in
    the incidence of lung tumours in female A/J mice (40 mice per chemical
    tested) when administered by gavage at doses of 10 mg/kg of body
    weight, 3 times weekly for 8 weeks (Bull & Robinson, 1985). Mice were
    started on treatment at 10 weeks of age and sacrificed at 9 months of
    age. No significant effects were observed with DBAN and DCAN. The
    differences in tumorigenic response are too small for meaningful
    rankings of the compounds for potency in this lung tumour-susceptible
    strain.

         DCAN was found to induce aneuploidy in  Drosophila (Osgood &
    Sterling, 1991) at a concentration of 8.6 mg/litre. On the other hand,
    DBAN produced inconsistent results but was tested at much lower
    concentrations (0.3 mg/litre) because of its higher degree of
    toxicity. Low levels of sodium cyanide (0.2 mg/litre) were also found
    to be active in this test system. Since DCAN is metabolized to
    cyanide, the authors suggested that the cyanide ion (CN-) was
    responsible for the response. DCAN is somewhat more efficiently
    converted to cyanide in rats than is DBAN (Pereira et al., 1984). This
    difference would be multiplied by the much higher concentration of
    DCAN that was tested.

         Daniel et al. (1986) found that the HANs were direct-acting
    electrophiles with the following decreasing order of reactivity: DBAN
    >> BCAN > CAN >> DCAN >> TCAN. The ability to induce DNA strand
    breakage in human CCRF-CEM cells was found to follow a considerably
    different order: TCAN >> BCAN > DBAN > DCAN > CAN. It is of
    interest that tumour-initiating activity paralleled alkylation
    potential in a cell-free system rather than an ability to induce
    strand breaks in DNA in intact cells or to induce mutation in
     Salmonella (BCAN > DCAN >> DBAN > TCAN = CAN = 0). If CAN is
    omitted from the group, mutagenicity parallels the extent to which the

    HAN is converted to cyanide  in vivo (Daniel et al., 1986). This
    discordant set of parallels suggests that some property may be
    affecting the ability of the test systems to measure the response or
    that the responses are not mediated through a common mechanism.
    Clearly, cytotoxic effects limited the responses of  Salmonella to
    the brominated HANs. Similar activity appeared to be affecting the
    mouse skin initiation/promotion studies, but only after a fairly
    robust response was observed. Some of the other effects may be only
    loosely associated with a health effect. For example, the induction of
    SSBs in DNA can arise from cytolethal effects. Moreover, such breaks
    reflect DNA repair processes as well as damage. Therefore, it is
    suggested that the carcinogenic potency of this class best parallels
    alkylation potential.

         TCAN was found to covalently bind with macromolecules in liver,
    kidney and stomach of the F344 rat (Lin et al, 1992). The covalent
    binding index was found to be the highest in the DNA of the stomach,
    followed by the liver, and was lowest in the kidney. The binding of
    14C was significantly higher when it was in the C2 position rather
    than in C1, indicating that the nitrile carbon is lost. The adducts
    formed were labile, and no specific adducts were identified. Adducts
    to blood proteins were also observed with TCAN. Covalent binding of
    DBAN or DCAN to DNA could not be demonstrated (Lin et al., 1986), but
    binding to proteins was apparently not investigated with these HANs.

         In conclusion, the HANs do possess carcinogenic and mutagenic
    properties in short-term tests. However, without appropriate long-term
    animal studies, the carcinogenic risk from HANs cannot be estimated.

    4.4.4  Comparative pharmacokinetics and metabolism

         The metabolism of the HANs has received some preliminary study,
    but little information exists on the pharmacokinetics of the parent
    compounds or their products. It is also important to note that some of
    the HANs inhibit enzymes that are important in the metabolism of other
    chemicals that are foreign to the body.

         Pereira et al. (1984) found the following percentage of the
    original doses of the HANs eliminated in the urine within 24 h as
    thiocyanate: CAN, 14%; BCAN, 12.8%; DCAN, 9.3%; DBAN, 7.7%; and TCAN,
    2.3%. This was compared with 42% of a dose of propionitrile. On the
    basis of this limited information and a general scheme for the
    elimination of cyanide from nitriles, published by Silver et al.
    (1982), Pereira et al. (1984) proposed that additional products of HAN
    metabolism would be as follows: CAN, formaldehyde; DHANs, formyl
    cyanide or formyl halide; and TCAN, phosgene or cyanoformyl chloride.
    These products would be direct-acting alkylating agents.

         Roby et al. (1986) studied the metabolism and excretion of DCAN
    labelled with 14C in either the C1 or C2 position in both male F344
    rats and B6C3F1 mice. The metabolic fate of the two carbons was
    significantly different in both mice and rats: C2 is metabolized much
    more efficiently to carbon dioxide, whereas a very much higher
    proportion of C1 is found as urinary metabolites, at least in mice.

    These results are consistent with the proposal of Pereira et al.
    (1984) suggesting metabolites that would be converted efficiently to
    carbon dioxide from C2.

         The HANs inhibit enzymes in the liver of the rat that are
    traditionally associated with the metabolism of foreign compounds.
    Pereira et al. (1984) demonstrated the inhibition of
    dimethylnitrosamine demethylase activity. This activity has been
    traditionally associated with the cytochrome P450 isoform 2E1,
    although there was no direct confirmation of this in the study.
    However, two forms of the enzyme activity were identified, one with a
     Km of 2  10-5 and the other with a  Km of 7  10-2. Based on a
    plot presented in the paper, DBAN appears to be inhibiting the
    high-affinity enzyme by either a non-competitive or uncompetitive
    mechanism. The kinetics of inhibition were examined  in vitro and the
    enzyme : inhibitor dissociation constants ( Kis) reported to be as
    follows: DBAN, 3  10-5; BCAN, 4  10-5; DCAN or TCAN, 2  10-4; and
    CAN, 9  10-2. Although not commented upon by the authors, the
    kinetics of inhibition by TCAN were clearly different from those of
    the DHANs, suggesting some differences in the mechanism or the form of
    the enzyme that might be affected. The authors examined the effects of
    DBAN or TCAN administered orally to rats at doses of 0.75 mmol/kg of
    body weight on the dimethylnitrosamine demethylase activity in the
    liver at 3 and 10 h after administration. TCAN significantly reduced
    the activity of the enzyme by about 30% at both time intervals, but
    DBAN did not, despite the fact that it was the more potent inhibitor
     in vitro. This could represent a difference in the extent to which
    the two compounds are absorbed systemically, or it could be related to
    the nature of the inhibition (i.e., reversible vs. irreversible).

         Ahmed et al. (1989) demonstrated inhibition of cytosolic GSTs by
    HANs  in vitro. Doses (mmol/litre) at which 50% inhibition of the
    activity of the enzyme GST occurred were as follows: DCAN, 2.49; TCAN,
    0.34; DBAN, 0.82; CAN, >10; and BAN, >10. This latter observation
    has not been established to occur in animals (Gao et al., 1996).
    Activation and inactivation of various DBPs are catalysed by various
    isoforms of GST (Pegram et al., 1997; Tong et al., 1998). In this
    case, toxicity is reduced by inhibition of GST; however, mutagenic
    activity of brominated THMs appears to depend upon a GSH pathway
    (Pegram et al., 1997). Similarly, GSTs appear to play an important
    role in the metabolism of HAAs (Tong et al., 1998).

         While these effects on the metabolism of other DBPs could be of
    importance, there are no data with which to relate these effects to
    concentrations that would be encountered in drinking-water.

    4.4.5  Mode of action

         The induction of skin tumours appears closely correlated with the
    alkylating potential within this class of chlorination by-products
    (Daniel et al., 1986). This suggests that the carcinogenic activity of
    the HANs may be related to mutagenic effects, despite the fact that
    cytotoxicity limits the capability of test systems to detect such
    activity. Cytotoxicity actually appears to inhibit the

    tumour-initiating activity responses rather than to amplify them, as
    has been seen with other DBPs. This suggests that the HANs retain some
    specificity for inducing cytotoxic responses in initiated cells.

         Another concern in this class is the ability of certain members
    to induce developmental delays. At present, interpretation of these
    results is somewhat clouded by the issues of interactions in the
    toxicity of the test compound with its vehicles, as discussed in the
    recent publication of Christ et al. (1996). No satisfactory
    explanation has been offered for these results. Such effects may be
    the result of fairly subtle changes in the pharmacokinetics and
    metabolism of the compound, or, as Christ et al. (1996) suggest, TCAN
    may have acted synergistically with a subthreshold effect of
    tricaprylin on developmental processes. This is suggested by a
    minimal, but consistent, response in treatment groups that received
    tricaprylin only relative to a group of naive controls.

         The potential importance of the HANs as cyanogens has not been
    extensively explored. As noted above, Pereira et al. (1984) indicated
    that significant portions of the dose of the HANs are eliminated as
    thiocyanate in the urine. Thus, cyanide release could be contributing
    significantly to the effects of these chemicals at high doses. 

    4.5  Halogenated hydroxyfuranone derivatives

         The halogenated hydroxyfuranones were first identified in the
    bleaching of pulp (Holmbom et al., 1984). The first member of this
    class, MX, was found because of its high mutagenic activity in
     Salmonella tester strains. In the same time frame, mutagenic
    activity had been associated with the chlorination of drinking-water
    (Meier, 1988). The high specific mutagenic activity of MX prompted
    examination of the possibility that it could contribute to the
    mutagenic activity that had been identified in chlorinated
    drinking-water. Subsequent experimentation confirmed this hypothesis.
    Estimates of the contribution of MX to the mutagenic activity in
    drinking-water ranged from 3% to 57% (Hemming et al., 1986; Meier et
    al., 1987a; Kronberg & Vartiainen, 1988). The lower estimates should
    probably be discounted, because the methods for recovering MX and
    other mutagens from water have varied. Moreover, it is important to
    recognize that these estimates apply only to MX itself. A variety of
    related compounds are produced that are also mutagens (Daniel et al.,
    1991b; DeMarini et al., 1995; Suzuki & Nakanishi, 1995), but which
    have received very little toxicological study. The contribution of
    these related compounds has not been estimated.

         The present section will focus on the research subsequent to that
    which identified MX as an important chlorination by-product. This
    review will confine itself to newer data that provide some perspective
    on possible mechanisms of action  in vivo. To avoid unnecessary
    redundancy, other furanones will be discussed only as they have been
    investigated to aid in an understanding of responses to MX.

    4.5.1  General toxicological properties and information on 
           dose-response in animals

         The lack of a commercial source of MX has limited research in
    experimental animals. However, a number of  in vivo studies and a
    carcinogenesis study of MX have been published (Bull et al., 1995;
    Komulainen et al., 1997). Reported epidemiological associations of
    drinking-water mutagenicity with cancer of the gastrointestinal and
    urinary tracts (Koivusalo et al., 1994b, 1996, 1997) provided
    additional impetus for investigating the compound.

         Three studies explicitly examined the acute toxicity of MX. The
    oral LD50 for MX in Swiss-Webster mice was determined to be 128 mg/kg
    of body weight when MX was administered for 2 consecutive days by
    gavage (Meier et al., 1987b). Doses of 70% of the LD50 and less
    (<90 mg/kg of body weight) had no significant effect on body weight
    of the mice, nor were they lethal. Most mice died within 24 h of
    receiving the first dose. Mice that died were found to have enlarged
    stomachs with moderate haemorrhagic areas in the forestomach. Very
    limited mortality was observed in weanling CD-1 mice administered a
    single dose of 144 mg/kg of body weight (Mullins & Proudlock, 1990).
    In this study, focal epithelial hyperplasia was observed in the
    stomach, and some vacuolation of the superficial villus epithelium was
    observed in the duodenum and jejunum. Evidence of increased numbers of
    mitotic figures was observed in the liver, and the possibility of some
    cytotoxicity was identified in the urinary bladder. Komulainen et al.
    (1994) administered MX in distilled water to male Wistar rats. Rats
    tolerated doses of 200 mg/kg of body weight administered by gavage 
    but displayed severe symptoms including dyspnoea, laborious breathing, 
    depressed motor activity and cyanosis at higher doses (Komulainen et
    al 1994). At necropsy, gastrointestinal inflammation was observed, 
    and oedema was noted in the lungs and kidneys. An LD50 of 230 mg/kg
    in 48 hours was identified in this study.

         The study of Meier et al. (1996) examined the effects of a 14-day
    course of MX treatment by gavage at a dose of 64 mg/kg of body weight
    per day on a number of enzyme activities in the liver of rats. MX
    treatment reduced hepatic levels of catalase, cytochrome P450
    reductase, aminopyrine demethylase and aromatic hydrocarbon
    hydroxylase. It did not affect fatty acyl CoA oxidase,
    glutamylcysteine synthetase, GST or glutathione peroxidase. The main
    result of such effects would be potential modifications of metabolism
    of various xenobiotics and endogenous biochemicals.

         A more extensive study of the effect of MX on enzyme activities
    in various tissues was conducted by Heiskanen et al. (1995) in Wistar
    rats. This study employed a constant daily dose of 30 mg/kg of body
    weight administered by gavage for 18 weeks as the low dose, whereas
    the higher dose was achieved by initiating treatment at 45 mg/kg of
    body weight (7 weeks) and raising it to 60 mg/kg of body weight
    (2 weeks) and to 75 mg/kg of body weight (5 weeks). A dose-related
    decrease in ethoxyresorufin- O-deethylase activity was observed in
    liver and kidney. MX appears to inhibit this enzyme's activity
    directly based upon  in vitro experiments conducted by the authors.
    However, the high concentrations required (0.9 mmol/litre) are
    unlikely to be approached systemically from the very low levels found

    in chlorinated drinking-water. The treatment also increased the
    activities of two phase 2 enzymes, uridine
    diphosphate-glucuronosyltransferase and GST, in the kidneys in a
    dose-dependent manner, but only in female rats. The health
    consequences of such modifications are uncertain, but could reflect an
    effect on physiological mechanisms associated with differences in sex.
    Again, it is important to recognize that these effects were produced
    at doses that were chronically, as well as acutely, toxic and unlikely
    to be remotely approached at the low concentrations found in
    chlorinated drinking-water.

         In a subchronic (14-18 weeks) toxicity study, Wistar rats (15 per
    sex per group) were given MX by gavage, 5 days per week, at doses of 0
    or 30 mg/kg of body weight (low dose) for 18 weeks or, in the
    high-dose group, at doses increasing from 45 to 75 mg/kg of body
    weight over 14 weeks. The high dose was finally lethal (two males and
    one female died) and caused hypersalivation, wheezing respiration,
    emaciation and tangled fur in animals. Increased water consumption,
    decreased body weights and food consumption, elevated plasma
    cholesterol and triglycerides, and increased urine excretion were
    noted in high-dose male rats. Urine specific gravity was decreased and
    the relative weights of the liver and kidneys were increased in both
    sexes at both doses in comparison with the controls. At both doses,
    duodenal hyperplasia occurred in males and females, and slight focal
    epithelial hyperplasia in the forestomach was observed in males.
    Splenic atrophy and haemosiderosis were seen in two high-dose females,
    and epithelial cell atypia was seen in the urinary bladder of one
    high-dose male and female. The frequency of bone marrow polychromatic
    erythrocytes with micronuclei was slightly increased only in low-dose
    male rats (Vaittinen et al., 1995).

    4.5.2  Toxicity in humans

         There have been no studies of the effects of these compounds in
    humans.

    4.5.3  Carcinogenicity and mutagenicity

    4.5.3.1  Studies in bacteria and mammalian cells  in vitro

         There have been extensive studies of the mutagenic activity of MX
    and related chemicals.  In vitro studies have extended knowledge
    beyond the initial characterization of simple mutagenic responses to
    (i) demonstrate effects in higher test systems, (ii) extend the data
    to other genotoxic end-points, (ii) characterize the mutagenic lesions
    produced in DNA and (iv) develop structural correlates. At higher
    levels of biological organization, a limited number of studies have
    been conducted to document that the genotoxic form of MX reaches the
    systemic circulation and to measure mutagenic effects in particular
    cell types  in vivo. 

         Meier et al. (1987b) demonstrated that MX induced chromosomal
    aberrations in CHO cells at concentrations as low as 4 g/ml  in 
     vitro. However, these authors were unable to demonstrate an

    increased frequency of micronuclei in the bone marrow of Swiss-Webster
    mice following two consecutive daily doses administered by gavage at
    70% of the LD50 (90 mg/kg of body weight  2). Treatment with
    30-300 mol of MX per litre (1 h) induced DNA damage in a
    concentration-dependent manner in suspensions of rat hepatocytes. DNA
    damage was induced in V79 Chinese hamster cells and in isolated rat
    testicular cells at the same concentrations as in hepatocytes. V79
    cells exposed to 2-5 mol of MX per litre (2 h) showed an increased
    frequency of SCE, whereas no significant effect on
    hypoxanthine-guanine phosphoribosyltransferase mutation induction was
    observed (Brunborg et al., 1991).

         Watanabe et al. (1994) found that the mutagenic activity of MX
    was effectively inhibited by sulfhydryl compounds such as cysteine,
    cysteamine, GSH, dithiothreitol and 2-mercaptoethanol. Pre-incubation
    of 0.5 g of MX with 15 g of cysteine in a phosphate buffer at 37C
    for 15 min prior to exposure of bacterial cells depleted the mutagenic
    activity of MX. Together with the result showing a change in the UV
    spectra, the authors suggested that sulfhydryl compounds inactivate MX
    by direct chemical interaction before MX induces DNA damage. On the
    other hand, a variety of antioxidants other than the sulfhydryl
    compounds showed no inhibitory effects. Investigation using structural
    analogues of cysteine revealed that the thiol moiety was indispensable
    for antimutagenic activity, and the amino moiety appeared to enhance
    the MX-inactivating reaction of the sulfhydryl group.

         Incubation of both rat and mouse hepatocytes with MX  in vitro 
    resulted in a dose-dependent increase in UDS at subcytotoxic
    concentrations (1-10 mol of MX per litre; 20-h incubation). Depletion
    of GSH stores by pretreatment of rat hepatocytes with buthionine
    sulfoximine did not result in a significant increase in UDS produced
    by MX. In contrast, MX did not induce UDS in mouse hepatocytes  ex 
     vivo either 3 or 16 h following administration of a single oral dose
    of 100 mg of MX per kg of body weight. Despite the ability of MX to
    produce repairable DNA damage, restricted access of MX to the liver
    may prevent a measurable UDS response  in vivo (Nunn et al., 1997).

         Jansson et al. (1995) examined MX and a related compound,
    3,4-(dichloro)-5-hydroxy-2(5H)-furanone (MA, also known as mucochloric
    acid), in the CHO hypoxanthine phosphoribosyl transferase  (hprt) 
    locus assay system where 6-thioguanine resistance (TGr) is the
    parameter measured. Both MX and MA induced TGr mutants. Indirect
    evidence was provided to suggest that the difference in sensitivity in
    the bacterial systems was related to differential ability to repair
    bulky adducts hypothesized to be induced by MX versus smaller adducts
    suggested to occur as a result of MA treatment. These results are
    interesting, considering the fact that Daniel et al. (1991b) found
    that MX and MA were of approximately equivalent potencies in inducing
    nuclear anomalies in gastrointestinal cells of the B6C3F1 mouse.

         Harrington-Brock et al. (1995) recently examined MX in the
    L5178/TK+1C3.7.2C mouse lymphoma system. A mutant frequency of 1027
    per 106 surviving cells was found. There was, however, a predominance
    of small-colony mutants. Small colonies more commonly arise from

    clastogenic effects than from point mutations. In parallel
    experiments, MX was found to have clastogenic effects (chromatid
    breaks and rearrangements). Point mutations and clastogenic effects
    were both observed at a medium concentration of 0.75 g/ml.

         MX and MA also induced micronuclei when applied to inflorescences
    of pollen mother cells of  Tradescantia (Helma et al., 1995). MX was
    approximately 5 times as potent as MA in this assay system.

         DeMarini et al. (1995) compared the mutation spectra induced by
    MX and extracts of water treated with chlorine, chloramine, ozone or
    ozone followed by chlorine or chloramine in  Salmonella typhimurium 
    strains TA98 and TA100. The mutation spectra induced in the  hisG46 
    codon displayed a predominance of the GAC mutation with MX, but there
    were also significant increases in the CTC and to a lesser extent ACC
    mutations. These latter mutations are not typical of MX. Since MX has
    never been identified as a by-product of ozonation, it is somewhat
    surprising that extracts of ozonated water produced a similar
    spectrum, even though these extracts are much less potent than those
    obtained from chlorinated or chloraminated water (i.e., they produced
    net increases in mutant colonies that are only about twice the
    spontaneous rate). The mutation spectra induced in the TA98
     hisD3052 allele more clearly differentiated between ozone and
    chlorinated or chloraminated water. In this case, virtually all of the
    frameshift mutations induced by raw and ozonated water extracts
    involved hotspot mutations, whereas only 30-50% of those induced by MX
    or extracts from water that had been treated with chlorine or
    chloramine involved the hotspot. Thus, the TA98 mutations are
    consistent with the hypothesis that the chemicals responsible for
    mutagenic effects of ozonated water are distinct from those induced by
    chlorination by-products.

         Hyttinen et al. (1996) found that MX and MA induce different
    mutation spectra in the DNA of  Salmonella typhimurium hisG46 codons
    (target codon sequence is CCC). The predominant mutation was GAC
    followed by ACC in MX-treated colonies, whereas CTC dominated the
    spectra produced by MA. MX primarily induced G:C -> T:A
    transversions, whereas MA produced G:C -> A:T transitions in the
    second base of the codon. The G:C -> T:A transversion in
     Salmonella was also observed in the  hprt gene of CHO cells
    (Hyttinen et al., 1996). Knasmuller et al. (1996) found the same
    difference in the mutation spectra of MX and MA in  S. typhimurium 
    mutants. These latter authors further found that
    3-chloro-4-(chloromethyl)-5-hydroxy-2(5H)-furanone (CMCF) produced the
    same transversion as MX, while chloromalonaldehyde produced the same
    transition found with MA.

         Some complex structure-activity relationships appear to occur
    with these two types of halofuranones. Kronberg et al. (1993) found
    that MA forms ethenocarbaldehyde derivatives with adenosine and
    cytidine, with chloroacetaldehyde being an intermediate. As pointed
    out by Knasmuller et al. (1996), ethenocytosine adducts formed by
    vinyl chloride cause G:C -> A:T transitions as reported for MA. The
    mutations induced by MX are similar to those produced by carcinogens

    that form bulky adducts, such as benzo [a]pyrene and 4-aminobiphenyl.
    Bulky adducts block replication leading to base substitutions
    according to the "adenine rule" (Strauss, 1991). As a consequence,
    there may be some significant differences in the health impact (e.g.,
    tumour site or character if they are found to be carcinogenic) of the
    mutagenic activities of these two chlorohydroxyfuranone derivatives.

         Ishiguro et al. (1988) examined structure-activity relationships
    for MX and related compounds. Their studies identified the chlorine
    substitution on C3 as being very important to the mutagenic activity
    of MX. Association of similar losses in mutagenic activity by removing
    the analogous chlorine from an open-ring structure compound,
    3-(dichloromethyl)-4,4-dichloro-2-chlorobutenoic acid, strongly
    supported this hypothesis.

         LaLonde and co-workers (LaLonde et al., 1991a,b, 1992; LaLonde &
    Xie, 1992, 1993) conducted a series of experimental and computational
    studies to relate the electronic structure of MX and related compounds
    with mutagenic activity within the class. Substitutions for the
    hydroxyl group led to reduction of mutagenic activity by a factor of
    100. Removal of the C3 or C6 chlorines from the structure reduced
    mutagenic activity by a factor of 10. Removal of the second chlorine
    at C3 caused a very large further reduction of mutagenic activity (by
    a factor of 1000) (LaLonde et al., 1991b). The mutagenicity appears to
    depend on the electron density at C2, C3 or C4 based upon 13C
    chemical shifts observed by nuclear magnetic resonance. Expansion of
    these studies supported the hypothesis that the mutagenic properties
    of the class paralleled the electrophilic character of chemicals
    within the class and the ability to stabilize a radical anion
    following acceptance of a single electron.

         Another difference in the chemistry of MA and MX, which might
    have biological implications, is that GSH readily displaces the
    chlorine on C4 of MA, greatly reducing its electrophilicity. On the
    other hand, GSH or  N-acetylcysteine reacts with MX to produce
    mixtures that are intractable to analysis, with the release of
    hydrogen sulfide (LaLonde & Xie, 1993).

    4.5.3.2  Studies in experimental animals

          In vivo studies are of two types: those that use bacterial
    systems to document absorption of MX (or mutagenic metabolite), and
    those in which effects on the test animal are directly measured.

         Fekadu et al. (1994) injected mixtures of repair-competent and
    repair-deficient  Escherichia coli K-12 cells intravenously into mice
    as test cells, and the animals were subsequently treated with 200 mg
    of test chemical per kg of body weight. Two hours later, the mice were
    sacrificed and cells recovered from various organs. MX, CMCF and MA
    were the test chemicals. The differential survival of the DNA
    repair-deficient strain versus a repair-competent variant is used to
    detect mutagenic activity. All three compounds significantly reduced
    recovery of the repair-deficient strain in the stomach, lung,
    intestine, liver, kidney and spleen. In a further experiment, the

    effects of lower doses of MX (4.3, 13 and 40 mg/kg of body weight)
    were investigated. Significantly depressed recovery was seen with MX
    doses as low as 4.3 mg/kg of body weight. MA did not modify recovery
    of the repair-deficient strain at doses less than or equal to 40 mg/kg
    of body weight. These data suggest that significant amounts of MX or a
    mutagenic metabolite reach the systematic circulation and at least
    reach the extracellular water. They do not clearly demonstrate effects
    in the target tissue of the experimental animal.

         Meier et al. (1996) found that only 0.3% of the original dose was
    excreted in a genotoxically active form in the urine of rats
    administered MX at a dose of 64 mg/kg of body weight for 14 days by
    gavage. No evidence of micronuclei induction was detected in
    peripheral blood erythrocytes in mice treated with a similar protocol.
    Whereas mutagenic activity was observed in urine at doses of 64 mg/kg
    of body weight, no significant mutagenic activity was observed at
    doses of 32 mg/kg of body weight and below.

         Brunborg et al. (1990, 1991) studied DNA damage induced by MX and
    other compounds in organs of rats using the alkaline elution assay (to
    detect strand breaks). While clear evidence of strand breaks was
    obtained with dibromochloropropane and
    2-amino-3,4-dimethylimidazo[4,5- f]quinoline, no significant effects
    were observed with MX after an intraperitoneal dose of 18 mg/kg of
    body weight or at oral doses of up to 125 mg/kg of body weight. The
    organs examined included the small and large intestine, stomach,
    liver, kidney, lung, bone marrow, urinary bladder and testis.

         Nishikawa et al. (1994) investigated cell proliferation and lipid
    peroxidation in the glandular stomach mucosa in Wistar rats given 0,
    6.25, 12.5, 25 or 50 mg of MX per litre in their drinking-water for
    5 weeks. Statistically significant cell proliferation increased in a
    dose-dependent manner up to 25 mg/litre. The MX treatment was also
    associated with increased lipid peroxidation levels in the gastric
    mucosa as well as in the urine, with loose dose dependence, although
    not at 50 mg/litre. Histopathologically, gastric erosion was noted in
    rats receiving 25 mg of MX per litre or more. These results suggest
    that MX may exert a gastric tumour-promoting action in rats, even at
    low doses that do not give rise to toxic effects, because of the clear
    dose-response relationship evident at low levels.

         The peripheral lymphocytes of male and female Han:Wistar rats
    exposed to MX at 30 or 45-75 mg/kg of body weight per day by gavage, 5
    days a week for 14-18 weeks, showed significant dose-related increases
    in SCEs at both levels of exposure in both sexes (Jansson et al.,
    1993).

         The peripheral lymphocytes of male Han:Wistar rats exposed to MX
    (25-150 mg/kg of body weight) by gavage on 3 consecutive days showed a
    significant dose-related increase in chromosomal damage measured as
    micronuclei, in addition to SCEs. Moreover, MX produced a significant
    dose-related increase in SCEs in the kidney cells of the exposed rats.
    However, the magnitude of the genotoxic responses observed was
    relatively weak (Maki-Paakkanen & Jansson, 1995).

         Daniel et al. (1991b) found that MX and MA induced nuclear
    anomalies in the epithelial cells of the gastrointestinal tract of
    B6C3F1 mice. Doses of 0.37 mmol/kg of body weight (approximately 80
    mg/kg of body weight) produced a modest increase in nuclear anomalies
    in the duodenum. There was no effect at 0.28 mmol/kg of body weight.
    Mullins & Proudlock (1990) and Proudlock & Crouch (1990) also found
    insignificant increases in nuclear anomalies in the non-glandular
    stomach, urinary bladder, jejunum and ileum. These latter authors
    noted that at the top dose at which nuclear anomalies were observed
    (144 mg/kg of body weight), there was significant irritation,
    inflammation and evidence of apoptotic cells in the gastrointestinal
    tract. These changes render the significance of the observed nuclear
    anomalies uncertain.

         MX was administered to Wistar rats (50 per sex per group) in
    drinking-water for 104 weeks at 0, 0.4, 1.3 or 5.0 mg/kg of body
    weight per day for males and 0, 0.6, 1.9 or 6.6 mg/kg of body weight
    per day for females. Dose-dependent increases in the incidence of some
    tumours were observed in rats, while the same MX doses had no obvious
    toxic effects on animals. Increases in tumours of the lung, mammary
    gland, haematopoietic system, liver, pancreas, adrenal gland and
    thyroid were observed, but few showed a clear dose-response (Table 21)
    (Komulainen et al., 1997).

    4.5.4  Comparative pharmacokinetics and metabolism

         There are very few data on the metabolism and pharmacokinetics of
    MX or related compounds. Ringhand et al. (1989) examined the
    distribution of radioactivity derived from 3-14C-MX in male F344
    rats. Approximately 35% of the radiolabel was eliminated in the urine
    and 47% in the faeces, with about 6% remaining in the body after 48 h.
    Neither the parent compound nor any specific metabolites were
    identified in any body compartment or fluid.

         Horth et al. (1991) studied the disposition of 3-14C-MX in male
    CD-1 mice. 14C was rapidly absorbed, reaching peak values in blood
    within 15 min of its administration. There was some evidence for
    binding to protein and retention of label within tissues, but no
    attempt was made to identify the chemical form in which the 14C was
    bound, so it was not clear whether MX was binding by virtue of its
    electrophilic character or whether this represented metabolic
    incorporation of metabolites of MX. Approximately 57% of the
    radioactivity was eliminated in the urine and 28% in the faeces. Less
    than 1% of the initial dose was retained in the carcass 120 h after
    administration, but most of this was associated with the stomach. It
    was stated that the urinary metabolites were polar, but no specific
    identifications were made.

         Komulainen et al. (1992) evaluated the pharmacokinetics of MX
    after a single oral or intravenous administration in Han:Wistar rats
    using 14C-labelled compound. Approximately 20-35% of the dose was
    absorbed into circulation from the gastrointestinal tract. The mean
    elimination half-life of the radioactivity in blood was 3.8 h. Traces
    of radioactivity remained in the blood for several days. The tissues

    lining the gastrointestinal and urinary tracts, kidney, stomach, small
    intestine and urinary bladder contained the highest radioactivity. The
    activity declined most slowly in the kidneys. Urine was the main
    excretion route, with 77% of the total radioactivity appearing in
    urine in 12 h and 90% in 24 h. No radioactivity was exhaled in air.
    After an intravenous administration of 14C-MX, the mean elimination
    half-life was much longer, 22.9 h, and the total elimination half-life
    was 42.1 h. Results indicate that MX is absorbed from the
    gastrointestinal tract to a considerable degree and is excreted in
    urine very rapidly. A fraction of MX or its metabolites is retained in
    blood for a longer period of time.

         No data are available in the scientific literature on the
    metabolism of MX or related compounds in humans.

         In conclusion, there are data to suggest that MX or a
    mutagenically active metabolite reaches the systemic circulation in
    experimental animals. Mutagenic activity has been detected in various
    organs and tissues using doses as low as 4.3 mg/kg of body weight
    (Fekadu et al., 1994). If these data are to have application to
    estimation of the hazards that MX presents to humans consuming
    chlorinated drinking-water, it is essential to understand whether MX
    or a metabolite reaches critical targets in the human body. An
    essential component of the information required would be an
    understanding of the metabolism and pharmacokinetics of MX and those
    of critical metabolites. The available data are too limited to provide
    much more than very general guidance in this area.

    4.6  Chlorite

    4.6.1  General toxicological properties and information on 
           dose-response in animals

         Concerns over chlorite in drinking-water first arose as chlorine
    dioxide began to play a role in the primary disinfection of
    drinking-water. Chlorite is the principal by-product of oxidative
    reactions of chlorine dioxide, but acidification of chlorite solutions
    is one method for generating chlorine dioxide for purposes of water
    disinfection (Aieta & Berg, 1986).

         Unless otherwise noted, references to chlorite will generally be
    to the sodium salt. This is the form most frequently studied. The term
    chlorite will be used if the authors expressed their doses in terms of
    chlorite; if dose levels were expressed as sodium chlorite, they will
    be identified as such. There is no reason to suspect that other salts
    of chlorite would exert inherently different toxicological effects, so
    this convention should not lead to confusion. A preparation of sodium
    chlorite with lactic acid is specifically excluded from consideration
    here because the actual composition of the product has not been
    specified (Scatina et al., 1984). Consequently, it is not clear that
    the data on this product have any relevance to chlorite or vice versa.


        Table 21. Summary of primary tumours observed in selected tissues in male rats after exposure to 
              3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) in drinking-water for 104 weeksa

                                                                                                          

    Tissue                                       Control     MX (mg/kg of body weight per day)   Pb
                                                                                              
                                                             0.4         1.3         5.0
                                                                                                          

    Integumentary system
       Skin, subcutaneous tissuec                50          50          50          50
          Basal cell tumourd                     1 (2%)                  1 (2%)      3 (6%)      0.0314
       Mammary glandsc                           50          50          50          49
          Adenocarcinomad                                                1 (2%)                  0.3162
          Fibroadenomad                                      1 (2%)      3 (6%)      1 (2%)      0.2996
          Fibromad                                           1 (2%)                              0.4749

    Respiratory system
       Lungsc                                    50          50          50          50
          Alveolar & bronchiolar carcinomasd     1 (2%)
          Alveolar & bronchiolar adenomasd       2 (4%)      1 (2%)      1 (2%)      7 (14%)     0.0015

    Haematopoietic system
       Multiple tissuesc                         50          50          50          50
          Lymphoma & leukaemiad                              3 (6%)      4 (8%)      3 (6%)      0.1527

    Digestive system
       Liverc                                    50          50          50          50
          Carcinomad                                                     2 (4%)      1 (2%)      0.1605
          Hepatocholangiocarcinomad                          1 (2%)                              0.4897
          Cholangiomad                                                   1 (2%)      4 (8%)      0.0009
          Adenomad                                           1 (2%)      2 (4%)      4 (8%)      0.0142
       Pancreasc                                 50          50          50          50
          Langerhans' cell carcinomad            4 (8%)      3 (6%)      5 (10%)     4 (8%)      0.3769
          Langerhans' cell adenomad              5 (10%)     8 (16%)     8 (16%)     12 (24%)    0.0116
          Acinar cell adenomad                   2 (4%)      3 (6%)      2 (4%)      4 (8%)      0.1243
                                                                                                          

    Table 21. (continued)

                                                                                                          

    Tissue                                       Control     MX (mg/kg of body weight per day)   Pb
                                                                                              
                                                             0.4         1.3         5.0
                                                                                                          

    Endocrine system
       Adrenal glandsc                           50          50          50          50
          Pheochromocytoma, malignantd                                   1 (2%)                  0.3213
          Pheochromocytoma, benignd              5 (10%)     2 (4%)      9 (18%)     3 (6%)      0.4830
          Cortical carcinomad                    2 (4%)      1 (2%)                              0.9447
          Cortical adenomad                      5 (10%)     2 (4%)      7 (14%)     14 (28%)    0.0001
       Thyroid glandsc                           49          50          50          50
          Follicular carcinomad                              1 (2%)      9 (18%)     27 (55%)    0.0000
          Follicular adenomad                    2 (4%)      20 (40%)    34 (68%)    21 (43%)    0.0045
          C-cell carcinomad                                              2 (4%)                  0.4478
          C-cell adenomad                        11 (22%)    7 (14%)     10 (20%)    11 (22%)    0.2459
                                                                                                          

    a  From Komulainen et al. (1997).
    b   P value from the one-sided trend test. A statistically positive trend at  P < 0.05 or lower.
    c  Values (reading across) = number of animals analysed.
    d  Values (reading across) = number of animals with one or more indicated tumours (frequency of animals 
       with tumour as percentage of examined animals.
    

         Early investigations of chlorite's toxic properties focused
    almost entirely on it potential ability to produce methaemoglobin and
    haemolysis. Heffernan et al. (1979a) examined the ability of sodium
    chlorite to induce methaemoglobinaemia in cats and Sprague-Dawley
    rats. When administered to cats as an oral bolus, as little as 20 mg
    of sodium chlorite per kg of body weight resulted in formation of
    significant amounts of methaemoglobin. Intraperitoneal doses of
    20 mg/kg of body weight in rats also induced methaemoglobin formation.
    However, when administered in drinking-water, no significant elevation
    in methaemoglobin was observed in cats (up to 1000 mg/litre as sodium
    chlorite) or rats (up to 500 mg/litre). Thus, chlorite must enter into
    the systemic circulation at a rapid rate, i.e., as a bolus dose, to
    induce methaemoglobin formation.

         Short-term toxic effects of chlorite were more systematically
    assessed in rats using gavage doses ranging from 25 to 200 mg/kg of
    body weight (Harrington et al., 1995a). Minor effects were observed at
    25 and 50 mg/kg of body weight. At 100 mg/kg of body weight and above,
    signs of haemolytic anaemia became apparent, with decreases in red
    blood cell count, haemoglobin concentration and haematocrit. These
    data support the idea that bolus doses were necessary to determine
    substantial effects on oxidative stress.

         Treatment of both cats and rats with sodium chlorite in
    drinking-water for extended periods (up to 90 days) resulted in
    decreases in red blood cell counts, haemoglobin concentrations and
    packed cell volume. These effects were observed with 500 mg of sodium
    chlorite per litre in cats (equivalent to 7 mg/kg of body weight per
    day) and with as little as 100 mg/litre in rats (equivalent to 10
    mg/kg of body weight per day) (Heffernan et al., 1979a). The changes
    in these blood parameters appeared to generally decrease in severity
    as the treatment was extended from 30 to 90 days, suggesting that
    adaptation to the treatment was occurring. In rats, however, red blood
    cell glutathione concentrations remained significantly depressed and
    2,3-diphosphoglycerate levels elevated through 90 days of treatment
    with concentrations of sodium chlorite as low as 50 mg/litre (5 mg/kg
    of body weight per day). In the cat, increased turnover of
    erythrocytes was detectable at concentrations of 100 mg/litre, with no
    significant effect being observed at 10 mg/litre. This latter
    concentration resulted in a daily dose of 0.6 mg/kg of body weight per
    day. Red blood cells drawn from rats treated with 100 mg/litre had
    significantly less ability to detoxify hydrogen peroxide that was
    generated by the addition of chlorite  in vitro. These data show that
    while the anaemia caused by haemolysis was largely compensated for in
    subchronic treatment of healthy rats, there was still evidence of
    oxidative stress being exerted by the chlorite treatment. Depletion of
    this reserve capacity could be of importance in individuals who are in
    a more compromised state (e.g., glucose-6-phosphate dehydrogenase
    deficiency) or who might be exposed to other haemolytic agents. The
    lowest concentration at which GSH was depleted significantly from
    control levels in rats was 50 mg/litre, and no effect was observed at
    10 mg/litre (equivalent to 1 mg/kg of body weight per day).

         The results of Heffernan et al. (1979a) have been generally
    confirmed by subsequent studies in a variety of species. Abdel-Rahman
    et al. (1980) and Couri & Abdel-Rahman (1980) obtained very similar
    effects in rats treated for up to 11 months. Moore & Calabrese (1980,
    1982) produced similar results in mice, and Bercz et al. (1982)
    demonstrated reduced red blood cell counts and decreased haemoglobin
    levels at similar doses in African green monkeys. These studies tended
    to identify altered forms of erythrocytes that are commonly associated
    with oxidative damage at treatment doses below those that produced
    actual anaemia (most consistently at a concentration of 100 mg/litre,
    with hints of such effects at lower doses).

         A more recent study employed doses of sodium chlorite
    administered by gavage to male and female Crl: CD (SD) BR rats (15 per
    sex per group) (Harrington et al., 1995a). Doses of 0, 10, 25 or 80 mg
    of sodium chlorite per kg of body weight per day were administered
    daily by gavage for 13 weeks (equivalent to 0, 7.4, 18.6 or 59.7 mg of
    chlorite per kg of body weight per day). This study is important
    because it included many of the standard parameters of subchronic
    toxicological studies, whereas previous studies had focused almost
    entirely on blood parameters. A gavage dose of 80 mg/kg of body weight
    per day produced death in a number of animals. It also resulted in
    morphological changes in erythrocytes and significant decreases in
    haemoglobin concentrations. Red blood cell counts were reduced
    slightly, but not significantly, at doses of 10 mg/kg of body weight
    per day in male rats, with further decreases being observed at 80
    mg/kg of body weight per day. Red blood cell counts were significantly
    depressed in female rats at doses of 25 mg/kg of body weight per day
    and above. As would be expected where haemolysis is occurring, splenic
    weights were increased. Adrenal weights were increased in females at
    25 and 80 mg/kg of body weight per day, whereas statistically
    significant changes were observed only at 80 mg/kg of body weight per
    day in males. Histopathological examination of necropsied tissues
    revealed squamous cell epithelial hyperplasia, hyperkeratosis,
    ulceration, chronic inflammation and oedema in the stomach of 7 out of
    15 males and 8 out of 15 females given 80 mg/kg of body weight per day
    doses. This effect was observed in only 2 out of 15 animals at the 25
    mg/kg of body weight per day dose and was not observed at all at 10
    mg/kg of body weight per day. Microscopic evaluations were made in 40
    additional tissues, and no treatment-related abnormalities were found.

         The Harrington et al. (1995a) study confirms the essential
    findings of previous studies and, in retrospect, justifies their focus
    on the blood cells as the critical target for chlorite toxicity. It
    also confirmed negative results of other studies that failed to
    identify significant effects in investigations of particular target
    organs (Moore et al., 1984; Connor et al., 1985).

    4.6.2  Reproductive and developmental toxicity

         Sodium chlorite did not exert any spermatotoxic effects in
    short-duration (1-5 days) tests (Linder et al., 1992).

         Moore et al. (1980b) reported that sodium chlorite administered
    at a concentration of 100 mg/litre throughout gestation and through
    28 days of lactation reduced the conception rate and the number of
    pups alive at weaning in A/J mice. A significantly reduced pup weight
    at weaning was interpreted as indicating that chlorite retarded growth
    rate.

         Groups of 4-13 Sprague-Dawley rats were treated on gestation days
    8-15 with sodium chlorite at concentrations of 0, 100, 500 or 2000
    mg/litre in drinking-water, by injection of 10, 20 or 50 mg/kg of body
    weight per day intraperitoneally or by gavaging with 200 mg/kg of body
    weight per day. Calculated daily doses of sodium chlorite administered
    to pregnant rats in drinking-water were 0, 34, 163 or 212 mg. Rats
    body weights were approximately 0.3 kg, giving estimated doses of 0,
    110, 540 or 710 mg/kg of body weight per day. Sodium chlorite at 20 or
    50 mg/kg of body weight per day intraperitoneally or at 200 mg/kg of
    body weight per day by gavage caused vaginal and urethral bleeding.
    Doses of 10, 20 and 50 mg/kg of body weight per day intraperitoneally
    caused 0%, 50% and 100% mortality of dams, respectively. No deaths
    were caused by sodium chlorite in the drinking-water, but the body
    weight and food consumption of the dams were decreased at 500 and 2000
    mg/litre. Blood smears from the dams injected intraperitoneally with
    all doses or drinking water containing 2000 mg of sodium chlorite per
    litre showed irregular, bizarre and ruptured erythrocytes. Injection
    of 10 or 20 mg/kg of body weight per day or drinking a solution
    containing 2000 mg/litre resulted in a decrease in litter size and an
    increase in stillbirths and resorption sites. Drinking 100 or 500 mg
    of sodium chlorite per litre did not produce any significant
    embryotoxicity. With all treatments, no significant gross soft tissue
    or skeletal malformations were observed. Postnatal growth of the pups
    was not affected by any treatment of the dams during the gestation
    period (Couri et al., 1982a,b).

         The effects of chlorite at 1 or 10 mg/litre in drinking-water for
    2.5 months prior to mating and throughout gestation were studied in
    Sprague-Dawley rats (Suh et al., 1983). This study indicated an
    increase in the incidence of anomalies in fetuses at both
    concentrations in two separate experiments; however, because the
    treatment groups were small (6-9 pregnant females per group), the
    effects were not considered statistically significant. Moreover, there
    were no consistent differences in either skeletal or soft tissue
    anomalies.

          Male and female Long-Evans rats were given 0, 1, 10 or 100 mg of
    sodium chlorite per litre of drinking-water. Males were exposed for 56
    days before mating and during 10 days of mating; females were treated
    for 14 days prior to mating, throughout the 10-day breeding period and
    gestation and through to day 21 of lactation. Males were evaluated for
    sperm parameters and reproductive tract histopathology following the
    breeding period. Dams and pups were necropsied at weaning. There was
    no effect on fertility, litter size or survival of neonates or on the
    weight of the testis, epididymis or cauda epididymis when males
    treated as described above were mated with these females. Decreases in
    the concentrations of triiodothyronine and thyroxine in blood were

    observed on postnatal days 21 and 40 in male and female pups exposed
    to 100 mg/litre. There were no effects at lower doses. Additionally,
    groups of males were exposed to 0, 10, 100 or 500 mg of sodium
    chlorite per litre for 72-76 days to confirm subtle observed changes
    in sperm count, morphology and movement. A significant increase in the
    percentage of abnormal sperm morphology and a decrease in the
    progressive sperm motility were observed for adult males at 100 and
    500 mg/litre (Carlton et al., 1987).

         Mobley et al. (1990) exposed groups of female Sprague-Dawley rats
    (12 per group) for 9 weeks to drinking-water containing 0, 20 or 40 mg
    of sodium chlorite per litre (0, 3 or 6 mg of chlorite per kg of body
    weight per day) beginning 10 days prior to breeding with untreated
    males and until the pups were sacrificed at 35-42 days
    post-conception. Animals exposed to a dose of 6 mg/kg of body weight
    per day exhibited a consistent and significant depression in
    exploratory behaviour on post-conception days 36-39. Exploratory
    activity was comparable between treated and control groups after
    post-conception day 39.

         In a two-generation study conducted by CMA (1997) and described
    in TERA (1998), Sprague-Dawley rats (30 per sex per dose) received
    drinking-water containing 0, 35, 70 or 300 mg of sodium chlorite per
    litre for 10 weeks and were then paired for mating. Males were exposed
    through mating, then sacrificed. Exposure for the females continued
    through mating, pregnancy, lactation and until necropsy following
    weaning of their litters. Twenty-five males and females from each of
    the first 25 litters to be weaned in a treatment group were chosen to
    produce the F1 generation. The F1 pups were continued on the same
    treatment regimen as their parents. At approximately 14 weeks of age,
    they were mated to produce the F2a generation. Because of a reduced
    number of litters in the 70 mg/litre F1-F2a generation, the F1
    animals were remated following weaning of the F2a generation to
    produce the F2b generation. Doses for the F0 animals were 0, 3.0,
    5.6 or 20.0 mg of chlorite per kg of body weight per day for males and
    0, 3.8, 7.5 or 28.6 mg of chlorite per kg of body weight per day for
    females. For the F1 animals, doses were 0, 2.9, 5.9 or 22.7 mg of
    chlorite per kg of body weight per day for males and 0, 3.8, 7.9 or
    28.6 mg of chlorite per kg of body weight per day for females. There
    were reductions in water consumption, food consumption and body weight
    gain in both sexes in all generations at various times throughout the
    experiment, primarily in the 70 and 300 mg/litre groups; these were
    attributed to a lack of palatability of the water. At 300 mg/litre,
    reduced pup survival, reduced body weight at birth and throughout
    lactation in F1 and F2, lower thymus and spleen weights in both
    generations, lowered incidence of pups exhibiting a normal righting
    reflex, delays in sexual development in males and females in F1 and
    F2, and lower red blood cell parameters in F1 were noted.
    Significant reductions in absolute and relative liver weights in F0
    females and F1 males and females, reduced absolute brain weights in
    F1 and F2, and a decrease in the maximum response to an auditory
    startle stimulus on postnatal day 24 but not at postnatal day 60 were
    noted in the 300 and 70 mg/litre groups. Minor changes in red blood
    cell parameters in the F1 generation were seen at 35 and 70 mg/litre,

    but these appear to be within normal ranges based on historical data.
    The NOAEL in this study was 35 mg/litre (2.9 mg/kg of body weight per
    day), based on lower auditory startle amplitude, decreased absolute
    brain weight in the F1 and F2 generations, and altered liver weights
    in two generations.

         Harrington et al. (1995b) examined the developmental toxicity of
    chlorite in New Zealand white rabbits. The rabbits (16 per group) were
    treated with 0, 200, 600 or 1200 mg of sodium chlorite per litre in
    their drinking-water (equal to 0, 10, 26 or 40 mg of chlorite per kg
    of body weight per day) from day 7 to day 19 of pregnancy. The animals
    were necropsied on day 28. There were no dose-related increases in
    defects identified. Minor skeletal anomalies were observed as the
    concentration of chlorite in water was increased and food consumption
    was depressed.

    4.6.3  Toxicity in humans

         The effects of chlorite have received some attention in
    toxicological and epidemiological investigations in human subjects.
    All of these studies were conducted at doses within an order of
    magnitude of concentrations of chlorite that might be expected in
    water supplies disinfected with chlorine dioxide. None pushed the
    limit of tolerance such that clear effects were observed. As a
    consequence, they are not informative for establishing a margin of
    safety.

         An experimental epidemiological study was conducted in the USA in
    a small city that had been using chlorine dioxide for some time in the
    summer months (April to October) to avoid taste and odour problems
    associated with the use of chlorine (Michael et al., 1981). Chlorine
    dioxide was generated from sodium chlorite that was mixed with
    chlorine gas and metered into the water. During the active use of
    chlorine dioxide, the chlorite concentrations in the water averaged
    5.2 mg/litre (range about 3-7 mg/litre). Subjects were monitored for
    11 parameters: haematocrit, haemoglobin, red cell count, white cell
    count, mean corpuscular volume, methaemoglobin, BUN, serum creatinine,
    total bilirubin, reticulocyte count and osmotic fragility of red blood
    cells. No effects could be associated with the switch of treatment
    from chlorine to chlorine dioxide disinfection. A total of 197 people
    were monitored in the exposed population, and there were 112
    non-exposed individuals. Each person served as his/her own control.

         Chlorine dioxide, free chlorine, chloramine and chlorate
    concentrations were also measured and were found to be 0.3-1.1,
    0.5-0.9, 0.9-1.8 and 0.3-1.8 mg/litre, respectively. The sampling for
    clinical measurements was done 1 week before chlorine dioxide
    disinfection began and 10 weeks into the cycle. Water samples taken
    during weeks 10-13 had chlorite levels that were systematically
    somewhat below those observed in the prior 9 weeks of sampling, and
    the same general trend was observable in other measures of chlorine
    dioxide and chlorate. This was not observed with chlorine or
    chloramines, suggesting that some change in water treatment had
    occurred. The authors provided no explanation for this change in water

    quality, but, since clinical samples were taken in week 10, this
    change in water quality could have resulted in lower exposure to
    chlorite and chlorate.

         The second set of evaluations of chlorite in humans involved
    direct administration of sodium chlorite in a rising-dose tolerance
    study and a follow-up study in which volunteers were treated for 12
    weeks, which were reported on in several publications. Lubbers et al.
    (1981, 1982) provided an overview of the studies. The detailed results
    of the rising-dose tolerance study were reported in Lubbers &
    Bianchine (1984), and those of the repeated-dose study in Lubbers et
    al. (1982, 1984a). A fourth paper (Lubbers et al., 1984b) reported
    results for three male volunteers that had glucose-6-phosphate
    dehydrogenase deficiency.

         The rising-dose tolerance study (Lubbers & Bianchine, 1984)
    involved administration of progressively increasing single doses of
    chlorite (0.01, 0.1, 0.5, 1.0, 1.8 or 2.4 mg/litre) in two 500-ml
    portions to a group of 10 healthy adult male volunteers. Doses were
    administered on days 1, 4, 7, 10, 13 and 16. In the interval between
    doses, clinical evaluations of the subjects were performed and a
    battery of clinical chemistry tests was performed on blood and urine
    samples. These latter tests were primarily directed at potential
    haematological effects of chlorite, but serum thyroxine and uptake of
    triiodothyronine were also determined. In addition, blood pressure,
    ECGs and other physiological parameters were monitored. No
    treatment-related effects were observed.

         In the repeated-dose study (Lubbers et al., 1984a), 10 male
    volunteers were administered 5 mg of chlorite per litre in a 500-ml
    portion for 12 weeks (0.036 mg/kg of body weight per day). Physical
    examinations and blood and urine analyses were conducted throughout
    the duration of treatment and for 8 weeks following the last dose of
    the solutions. None of the parameters investigated was found to fall
    outside the normal range; although there were some consistent changes
    in values with time, none of these appeared to be related to chlorite
    treatment.

         Three individuals with glucose-6-phosphate dehydrogenase
    deficiency were identified in the course of the study. This genetic
    disorder makes individuals more sensitive to oxidative damage, which
    is frequently manifested as increased methaemoglobin production and
    haemolysis when the individuals are exposed to oxidative chemicals in
    sufficient doses. All three individuals were treated with chlorite in
    the same concentrations and in the same manner as described for the
    study of normal individuals (Lubbers et al., 1984b). No clinically
    significant changes were found in these individuals.

         A study (Ames & Stratton, 1987) was conducted of renal dialysis
    patients in California (USA) after a water district introduced
    chlorine dioxide as a drinking-water disinfectant but failed to inform
    the clinic for 12 months. Water treatment at the clinic consisted of
    ion exchange, GAC, 5-m filtration and reverse osmosis. Chlorite

    levels measured after this treatment were 0.02-0.08 mg/litre, but
    there were periods during which no chlorine dioxide was added, and
    exposures to the patients may have been lower. Measures for 28 serum
    and haematological parameters were available for 17 renal dialysis
    patients for a period of 3 months before and 1 month after exposure.
    Methaemoglobin measures were not available. Only one measure was
    statistically associated with the use of water disinfected by chlorine
    dioxide: serum uric acid declined by 10% after exposure to disinfected
    water, a change that was not considered clinically important. The
    study found no evidence of anaemia or other adverse effects of
    chlorine dioxide-disinfected water for these renal dialysis patients,
    but the interpretation of these results is severely limited because of
    the small sample size and apparently very low exposures.

         Collectively, these studies suggest that humans are probably not
    sensitive to the concentrations of chlorite that are likely to be
    found in water disinfected with chlorine dioxide. Some safety factor
    is present in these data, because it is unlikely that concentrations
    of chlorite would exceed 1 mg/litre with new methods of application.
    However, these studies provide little information relative to the
    actual margin of safety that exists between those concentrations seen
    or administered and concentrations that would lead to clear adverse
    effects. Consequently, these studies do not imply that the
    concentrations of chlorite in drinking-water should be without limits.

    4.6.4  Carcinogenicity and mutagenicity

         Sodium chlorite was reported to produce a concentration-dependent
    increase in revertants in  Salmonella typhimurium strain TA100 in
    both the presence and absence of rat liver S9 fraction (Ishidate et
    al., 1984). A linear dose-response curve was observed, and the net
    number of revertants produced at 0.3 mg per plate was 88. The S9 mix
    used for metabolic activation was from the liver of F344 rats
    pretreated for 5 days with polychlorinated biphenyls at 500 mg/kg of
    body weight.

         Meier et al. (1985b) evaluated chlorite in the mouse micronucleus
    assay, the mouse bone marrow cytogenetics assay and the mouse sperm
    head abnormality assay. The doses administered were 0.2, 0.5 or 1 mg
    per mouse or approximately 40 mg/kg of body weight at the highest
    dose. No statistically significant results were found in any of the
    tests. In a later reference, it was indicated that chlorite also
    induced chromosomal aberrations (Kurokawa et al., 1986b), but the data
    were not provided. Hayashi et al. (1988, 1989) found an increase in
    micronuclei in the bone marrow of mice given 0, 7.5, 15, 30 or 60
    mg/kg of body weight by intraperitoneal injection at doses of 15 and
    30 mg/kg of body weight. In a repeat study in which mice were given 0
    or 15 mg/kg of body weight on 4 successive days, no increase in
    micronuclei was observed. In a study using the oral route with doses
    of 0, 37.5, 75, 150 or 300 mg/kg of body weight, a significant
    increase in micronuclei was observed only at 150 mg/kg of body weight.

         In a carcinogenicity study, sodium chlorite was administered to
    F344 rats (50 per sex per dose) at concentrations of 0, 300 or
    600 mg/litre of drinking-water (equivalent to 0, 18 or 32 and 0, 28 or
    41 mg of chlorite per kg of body weight per day in males and females,
    respectively) and to B6C3F1 mice (50 per sex per dose) at
    concentrations of 250 or 500 mg/litre (equivalent to 0, 36 or 71 mg/kg
    of body weight per day) for 85 weeks (Kurokawa et al., 1986b). The
    rats became infected with a Sendai virus in all groups, which resulted
    in the termination of the study after only 85 weeks. There was a
    statistically significant increase in the incidence of hyperplastic
    nodules in male mice treated with 250 mg/litre, but not in females.
    The incidence of these lesions did not increase when the dose of
    chlorite was increased to 500 mg/litre. Hepatocellular carcinomas were
    too few for their observation to add anything substantive to the
    evaluation. There were no other treatment-related changes in the
    incidence of other tumours in either male or female mice.

          Groups of 50 male and female B6C3F1 mice were given 0, 250 or
    500 mg of sodium chlorite per litre in the drinking-water for 80 weeks
    (0, 36 or 71 mg of chlorite per kg of body weight per day). A small,
    but statistically significant ( p < 0.05), increase in the incidence
    of lung adenomas was observed at 500 mg/litre. The authors noted that
    this was not accompanied by the appearance of lung adenocarcinomas and
    that the incidence was within the range of historical controls; thus,
    it was not possible to conclude from these data that chlorite induced
    lung tumours (Yokose et al., 1987).

         In an associated experiment, Kurokawa et al. (1984) assessed the
    ability of sodium chlorite to promote skin tumours in a group of
    20 female Sencar mice. These mice were initiated with a single topical
    application of 20 nmol (5.1 g) of dimethylbenzanthracene in acetone
    followed by 0.2-ml applications of sodium chlorite at 20 mg/ml in
    acetone twice weekly for 51 weeks. A group of 15 female mice given a
    single application of dimethylbenzanthracene followed by applications
    of acetone were used as controls. This treatment resulted in 5 of 25
    mice having squamous cell carcinomas at 52 weeks. No tumours were
    found in the corresponding initiated control mice. Both TPA and
    benzoyl peroxide produced increased tumour incidence in
    dimethylbenzanthracene-initiated mice. These data indicate the
    potential for a weak tumour-promoting activity for sodium chlorite.
    However, no dose-response information has been forthcoming in the
    literature.

    4.6.5  Comparative pharmacokinetics and metabolism

         Some limited data on the absorption, distribution and excretion
    of chlorite have been developed in rats using 36Cl-labelled chlorite.
    The label was absorbed with a half-life of about 3.5 min and
    eliminated with a terminal half-life of 35.2 h (Abdel-Rahman et al.,
    1982b, 1984b). In 72 h, approximately 35% of the label was recovered
    in the urine and another 5% in the faeces. In the urine, 32% of the
    administered dose was determined to be chloride, whereas 6% was found
    to be chlorite, utilizing a fractionation procedure developed in a
    prior study (Abdel-Rahman et al., 1980). While these studies did not

    determine the form of the radiolabel found in blood, plasma and
    tissues, it was clear that there were significant differences in the
    behaviour of the label derived from chlorite and chlorate. However,
    the efforts in this area have been seriously hampered by the lack of
    an analytical method to discriminate between chlorine dioxide,
    chlorite, chlorate and chloride  in vivo. 

    4.6.6  Mode of action

         The adverse effects of chlorite appear to be mediated through its
    activity as an oxidant. However, this question has received very
    limited attention, except for the involvement of oxidation in its
    haematological effects. Heffernan et al. (1979b) demonstrated that
    chlorite was consumed during the oxidation of haemoglobin to
    methaemoglobin  in vitro. It was also observed that, unlike
    methaemoglobin induction by nitrite, the action of chlorite also
    depleted the red blood cells of GSH, and that this could be partially
    counteracted by including glucose in the incubation medium. The
    oxidative action of chlorite could be associated with the production
    of hydrogen peroxide as measured by the formation of complex I with
    catalase. This production of hydrogen peroxide was associated with
    oxidative damage by demonstrating that it could also be attenuated by
    the inclusion of glucose in the medium. In a dose-response comparison,
    it could be demonstrated that the loss of GSH and the loss of catalase
    activity paralleled one another and occurred at concentrations an
    order of magnitude lower than those required for methaemoglobin
    formation. This is consistent with the behaviour of other oxidants
    that produce haemolytic anaemia. These observations also appear to
    explain why destruction of the red blood cell (measured as decreased
    haematocrit, decreased haemoglobin concentrations and increased red
    blood cell turnover) is a much more sensitive and important measure of
    chlorite toxicity than methaemoglobin formation.

    4.7  Chlorate

    4.7.1   General toxicological properties and information on 
            dose-response in animals 

         Toxicological data on chlorate in the open scientific literature
    are limited to two short-term studies in dogs (Sheahan et al., 1971;
    Heywood et al., 1972) and a series of studies that focused primarily
    on its ability to induce oxidative damage in the blood of rats and
    chickens (Abdel-Rahman et al., 1980; Couri & Abdel-Rahman, 1980) and
    African green monkeys (Bercz et al., 1982). Two short-term studies,
    one in dogs and one in rats, carried out by Bio/Dynamics Inc. in 1987,
    were reviewed by WHO (1996). In the dog study, a NOAEL of 360 mg/kg of
    body weight was identified based on no significant effects on any
    measured parameter. In the rat study, a NOAEL of 100 mg/kg of body
    weight was identified based on haematological effects at the highest
    dose (1000 mg/kg of body weight). There is also a single subchronic
    study of toxicity conducted in rats in which histopathological
    examination of tissues was performed (McCauley et al., 1995). This
    limited data set has hindered attempts to establish a guideline value
    for chlorate in drinking-water (WHO, 1993).

         The studies in dogs documented the fact that high acute doses of
    1 or 2 g/kg of body weight induce methaemoglobinaemia (Sheahan et al.,
    1971; Heywood et al., 1972). In addition, Heywood et al. (1972)
    administered lower doses (200-300 mg/kg of body weight) for 5 days and
    observed no clinical signs at doses lower than 300 mg/kg of body
    weight per day. Those doses that produced some evidence of
    methaemoglobinaemia were also found to have produced some
    discoloration of the kidneys and haematogenous cases in renal tubules
    at necropsy.

         No consistent effects were observed when chlorate was
    administered to rats at concentrations of 10 or 100 mg/litre
    (equivalent to 1 or 10 mg/kg of body weight per day) for 12 months
    (Couri & Abdel-Rahman, 1980). These authors documented some loss of
    the normal sensitivity of erythrocytes to osmotic shock at these
    doses, however.

         Bercz et al. (1982) administered drinking-water containing sodium
    chlorate to African green monkeys for a total of 8 weeks in a
    rising-dose experiment. Drinking-water concentrations were 25, 50,
    100, 200 or 400 mg/litre, equivalent to 4, 7.5, 15, 30 or 58.4 mg/kg
    of body weight per day. Chlorate was found to be without significant
    effect on a number of serum parameters related to oxidative damage and
    thyroid hormone levels at concentrations of up to 400 mg/litre.

         In the subchronic study (McCauley et al., 1995), concentrations
    of chlorate of 3, 12 or 48 mmol/litre in drinking-water were provided
    to both male and female Sprague-Dawley rats for 90 days. These
    concentrations correspond to 250, 1000 and 4000 mg of chlorate per
    litre, equal to 30, 100 or 510 mg/kg of body weight per day in males
    and 42, 164 or 800 mg/kg of body weight per day in females, based on
    measured water consumption of each group. Body weight gain was sharply
    curtailed in both sexes at the highest concentration. These effects
    were generally paralleled by smaller organ weights (except for brain
    and testes). Some decreases in haemoglobin, haematocrit and red blood
    cell counts were observed at this same dose. Pituitary lesions
    (vacuolization in the cytoplasm of the pars distalis) and thyroid
    gland colloid depletion were observed in both the mid- and high-dose
    groups of both sexes. The NOAEL in this study was 30 mg/kg of body
    weight per day.

    4.7.2  Reproductive and developmental toxicity

         Suh et al. (1983) examined the effects on fetal development of
    chlorate at 0, 1 or 10 mg/litre administered to rats for 2.5 months
    prior to mating and throughout gestation. This was a very limited
    study, involving only six female rats per treatment group. Therefore,
    the apparent increase of anomalous fetuses from 30.7% in the control
    group to 52% and 55.2% in the groups receiving 1 and 10 mg/litre,
    respectively, was not statistically significant. The abnormalities
    were limited to relatively mild skeletal defects (missing sternebra
    and rudimentary ribs).

         A teratogenicity study carried out by Bio/Dynamics Inc. in 1987
    was reviewed by WHO (1996). In this study, rats given 0, 10, 100 or
    1000 mg of chlorate per kg of body weight per day on days 6-15 of
    gestation showed no effects on maternal or fetal health.

    4.7.3  Toxicity in humans

         There have been sporadic reports of poisoning with sodium or
    potassium salts of chlorate (Temperman & Maes, 1966; Mengele et al.,
    1969; Yoshida et al., 1977; Bloxham et al., 1979; Helliwell & Nunn,
    1979; Steffen & Seitz, 1981). Most of these cases involved ingestion
    of preparations of sodium chlorate used for pesticidal purposes. The
    symptomatology observed is consistent with that observed in the acute
    studies in dogs identified above. There was generally evidence of
    oxidative damage to erythrocytes, methaemoglobin formation and the
    renal complications of haemolytic anaemia. The lethal dose to humans
    has been estimated to be in the range 20-30 g.

         A study in 10 male human volunteers was conducted by
    administering solutions of 0.01-2.4 mg of chlorate per litre in two
    500-ml portions (highest dose 0.034 mg/kg of body weight per day) in a
    6-day rising-dose tolerance design (Lubbers & Bianchine, 1984). No
    adverse effects were noted. This test of acute studies was followed by
    an experiment (Lubbers et al., 1981) that provided 500 ml of water
    containing 5 mg of chlorate per litre per day to 10 subjects for
    12 weeks (average dose 0.036 mg/kg of body weight per day). Volunteers
    in both studies were monitored using a battery of clinical and
    physiological parameters and routine physical examinations throughout
    the course of the study and for 8 weeks following termination of
    treatments. Again, no adverse effects were observed.

    4.7.4  Carcinogenicity and mutagenicity

         There are no published studies of the carcinogenic potential of
    chlorate administered alone. Sodium and potassium chlorate were
    evaluated as promoters of renal tumours in
     N-ethyl- N-hydroxyethylnitrosamine (EHEN)-initiated F344 rats.
    Sodium chlorate and potassium chlorate were administered in the
    drinking-water for 28 weeks. There was an increased incidence of renal
    cell tumours (7/15 rats) in the EHEN-initiated group treated with
    sodium chlorate, but no effect was observed with potassium chlorate
    (1/5) relative to control rats (2/15). The small numbers of animals
    used in this study make the treatment groups indistinguishable from
    one another statistically (Kurokawa et al., 1985b).

         Chlorate has long been known to select nitrate
    reductase-deficient mutants of  Aspergillus nidulans (Cove, 1976).
    However, Prieto & Fernandez (1993) demonstrated that there is also a
    mutagenic effect of chlorate in  Chlamydomonas reinhardtii and
     Rhodobacter capsulatus. Chlorate failed to induce mutations in the
    BA-13 strain of  Salmonella typhimurium. The positive mutagenic
    effects were separated from simple selection of nitrate reductase
    mutants by incubating cells in nitrogen-free media. Lack of nitrogen
    prevents cell division during the treatment period. In the case of

     C. reinhardtii, significant increases in mutants were observed at
    concentration of 4-5 mmol/litre and above.

         Meier et al. (1985b) examined chlorate in assays for micronuclei
    and chromosomal aberrations in bone marrow and sperm head anomalies,
    but all findings were negative.

    4.7.5  Mode of action

         Some research has been directed towards establishing the
    mechanisms by which chlorate oxidatively damages erythrocytes and
    their contents. There is a characteristic delay in the production of
    methaemoglobin by chlorate when erythrocytes are incubated in the
    presence of chlorate (Singelmann et al., 1984). It has been suggested
    that this delay was due to the conversion of chlorate to chlorite
    (Heubner & Jung, 1941; Koransky, 1952), but there is no direct
    evidence to support this view. A competing hypothesis suggested that
    chlorate formed a complex with methaemoglobin, which autocatalytically
    increased methaemoglobin formation. This suggestion is supported by
    experiments demonstrating that the formation of methaemoglobin
    accelerates the further formation of methaemoglobin in the presence of
    chlorate (Huebner & Jung, 1941; Jung, 1947, 1965). It is further
    supported by the observation that compounds that compete for binding
    of chlorate to methaemoglobin (e.g., azide or cyanide) block the
    effect.

         The properties of the erythrocyte membrane are also modified by
    chlorate. Increased resistance to haemolysis is the most readily
    observed effect (Singelmann et al., 1984). These effects appear to be
    related to the formation of high molecular weight complexes of
    erythrocytic proteins (Singelmann et al., 1984). These changes could
    not be reversed by disulfide reduction. The formation of these
    complexes was associated with the loss of activity of several enzymes,
    the most sensitive being glucose-6-phosphate dehydrogenase (Singelmann
    et al., 1984; Steffen & Wetzel, 1993). The inactivation of this enzyme
    accounts for the insensitivity of chlorate-induced methaemoglobinaemia
    to treatment with methylene blue. Reduction of nicotinamide adenine
    dinucleotide phosphate (NADP) by the pentose pathway is necessary for
    methylene blue to be effective. The cross-linking of protein is not
    limited to cytosolic proteins and methaemoglobin, because
    cross-linking of membrane proteins has also been demonstrated (Steffen
    & Wetzel, 1993). The cross-linking requires the presence of
    haemoglobin. Similar changes are induced by hypochlorite, but in this
    case haemoglobin is not necessary.

         The oxidative damage to the erythrocyte appears to be the basis
    of chlorate's renal toxicity. This hypothesis is supported primarily
    by the observation that species less sensitive to methaemoglobin
    formation are also resistant to the nephrotoxic effects of chlorate
    (Steffen & Wetzel, 1993). The observation is consistent with the
    finding of haematogenous casts in kidney tubules of dogs treated with
    doses of chlorate that induce methaemoglobin and their absence in dogs
    treated with slightly lower doses that did not produce
    methaemoglobinaemia (Heywood et al., 1972).

    4.8  Bromate

    4.8.1  General toxicological properties and information on 
           dose-response in animals 

         The acute toxic effects of bromate (administered as either the
    potassium or sodium salt) have been studied in F344 rats, B6C3F1 mice
    and Syrian golden hamsters (Kurokawa et al., 1990). The mean LD50
    values in these species ranged from 280 to 495 mg/kg of body weight,
    with slightly but consistently lower values found in males than in
    females of each species. Mice appear to be somewhat more sensitive
    than the other two species, but the lethal doses are remarkably
    similar across species. Toxic signs and symptoms at lethal doses
    included suppressed locomotor activity, ataxia, tachypnoea,
    hypothermia, diarrhoea, lacrimation and piloerection. Hyperaemia of
    the stomach and congestion of lungs were observed at autopsy. Damage
    to renal tubules was seen microscopically, including necrosis in the
    proximal tubular epithelium. Regenerative changes were observed from
    48 h to 2 weeks after treatment. These effects were less marked in
    mice and hamsters. No glomerular lesions were observed in any species.

         Treatment of rats for 10 weeks with potassium bromate
    concentrations of 250, 500, 1000, 2000 or 4000 mg/litre of
    drinking-water established a maximally tolerated concentration of less
    than 1000 mg/litre. As treatments were extended to 13 weeks, elevated
    levels of glutamate-oxalate transaminase (GOT), glutamate-pyruvate
    transaminase (GPT), LDH, AP and BUN were observed in blood samples
    (Onodera et al., 1986; Kurokawa et al., 1990).

         Eosinophilic droplets were observed in the cytoplasm of the
    proximal renal tubule cells in male F344 rats receiving 600 mg of
    potassium bromate per litre for 12 weeks (Onodera et al., 1986;
    Kurokawa et al., 1990). These droplets were determined to be
    eosinophilic bodies rather than hyaline droplets. Lipofuscin pigments
    were also observed in the proximal tubular epithelium.

         Dogs, rats and monkeys were fed bread or flour treated with up to
    200 mg of potassium bromate per kg of body weight for up to 17 months
    (FAO/WHO, 1989). These studies revealed no adverse effects, but, as
    pointed out by Kurokawa et al. (1990), substantial portions of bromate
    are presumed to be converted to bromide during the dough-making
    process. It is also important to note that the numbers of animals
    included in these studies were quite limited. Subsequent studies
    conducted for longer periods of time indicated an increase
    periarteritis in male rats and pathology to the adrenal glands in
    female rats (Fisher et al., 1979).

         Lifetime studies in female rats administered potassium bromate in
    drinking-water found significant increases in GPT, albumin/globulin
    ratios, serum potassium ion and cholinesterase activity at
    concentrations of 500 mg/litre. Slight increases in BUN were also
    observed at this dose (Kurokawa et al., 1990).

    4.8.2  Toxicity in humans

         Human poisonings have been associated with the ingestion of
    sodium bromate and potassium bromate. Many of these poisonings result
    from accidental or deliberate ingestion of preparations used as
    neutralizers in permanent wave kits (Warshaw et al., 1985; Lue et al.,
    1988). Clinical signs of bromate poisoning include anaemia and
    haemolysis, renal failure and hearing loss. Loss of hearing appears to
    be more common in adults than in children (Lichtenberg et al., 1989).
    The hearing loss and renal failure can have a prolonged course in
    some, but not all, people poisoned by bromate (Kuwahara et al., 1984).
    Poisoning with bromate is frequently fatal when doses exceed 6 g
    (Kurokawa et al., 1990).

    4.8.3  Carcinogenicity and mutagenicity

         IARC evaluated potassium bromate in 1986 and concluded that there
    is sufficient evidence for its carcinogenicity in experimental
    animals, whereas no data were available on its carcinogenicity to
    humans. On this basis, potassium bromate was assigned to Group 2B: the
    agent is possibly carcinogenic to humans (IARC, 1986, 1987).

         In 1992, the Joint FAO/WHO Expert Committee on Food Additives
    (JECFA) evaluated potassium bromate and concluded that it was
    genotoxic and carcinogenic. On this basis, JECFA concluded that the
    use of potassium bromate as a flour treatment agent was not
    appropriate (FAO/WHO, 1993).

         Potassium bromate was found to be weakly mutagenic in
     Salmonella typhimurium strain TA100 when incubated with rat S9
    fraction for metabolic activation (Kawachi et al., 1980; Ishidate et
    al., 1984). Negative results were found in strains TA98, TA1535,
    TA1537 and TA1538 (Kurokawa et al., 1990). Potassium bromate was also
    inactive in  Escherichia coli Wptry- and  E. coli WP2try- his-
    (Ishidate et al., 1984). Bromate was later found to be active in
     S. typhimurium strains TA102 and TA104, which were developed to
    detect compounds that generate oxygen radicals (Kurokawa et al.,
    1990).

         Potassium bromate also induced chromosomal aberrations in a
    Chinese hamster fibroblasts cell line. The concentrations required
    were, however, very high (>30 mmol/litre) (Ishidate et al., 1984).
    Such high doses may induce changes by indirect mechanisms.

         Bromate appears capable of inducing micronuclei  in vivo. 
    Significant increases in the frequency of micronuclei were observed in
    polychromatic erythrocytes when potassium bromate was administered by
    either the oral or intraperitoneal route (Hayashi et al., 1988;
    Nakajima et al., 1989). Positive results were obtained at doses of
    24 mg/kg of body weight when administered intraperitoneally. Oral
    doses of less than 100 mg/kg of body weight were negative. Ms/Ae and
    CD-1 mice were found to be equally sensitive to these effects.

         Several reports of bromate-induced cancer in experimental animals
    are available. The clearest evidence comes from studies in F344 rats
    (Kurokawa et al., 1983, 1986a, 1987a; DeAngelo et al., 1998). The
    dose-response curves for the principal target organs, the kidney and
    peritoneum, are provided in Figure 7. Tumours found in the kidney were
    of tubular origin, with significantly increased numbers of both
    adenomas and adenocarcinomas being observed in both males and females
    (Kurokawa et al., 1983). The peritoneal tumours were mesotheliomas,
    but treatment-related increases were observed only in male rats. In
    the Kurokawa et al. (1983) study, concentrations of potassium bromate
    of 0, 250 or 500 mg/litre correspond to doses of 0, 9.6 or 21.3 mg of
    bromate per kg of body weight per day in males and 0, 9.6 or 19.6
    mg/kg of body weight per day in females (as cited in IARC, 1986 and
    WHO, 1996). The Kurokawa et al. (1986a) study represented, in part, a
    repeat of the earlier study, except that more doses were included and
    only male rats were studied. Concentrations in this study were 0, 15,
    30, 60, 125, 250 or 500 mg of potassium bromate per litre,
    corresponding to 0, 0.7, 1.3, 2.5, 5.6, 12 or 33 mg of bromate per kg
    of body weight per day (as cited in WHO, 1996). The study provided
    verification of the ability of potassium bromate to induce both renal
    cell tumours and mesothelioma. An increased incidence of renal tumours
    was observed at 125 mg of potassium bromate per litre of
    drinking-water. Results with mesothelioma were not observed at a dose
    of 250 mg/litre, but this is partially attributable to the smaller
    number of animals per treatment group in this experiment. However, the
    incidences of mesothelioma were very similar at 500 mg/litre in the
    two studies. Significant increases in the occurrence of dysplastic
    foci of the kidney (considered to be preneoplastic lesions) were found
    in groups at doses higher than 30 mg/litre.

         The time course of renal tumour development was examined in a
    third study (dose-response provided in panel C of Figure 7) by
    Kurokawa et al. (1987a). These experiments were designed to include
    sacrifices of animals at 13, 26, 39 or 52 weeks as well as the
    104-week period examined in prior studies. Additional groups were
    included, however, that involved treatment for the above periods, but
    the animals were maintained on bromate-free water until 104 weeks
    before they were sacrificed. The concentration of potassium bromate
    was 500 mg/litre during active treatment periods (average dose 32.3 mg
    of bromate per kg of body weight per day). If animals were held for
    104 weeks, 13 weeks of treatment was sufficient to produce the same
    tumour incidence as was produced by longer treatment periods.
    Moreover, the same incidence of renal cell tumours was observed in
    animals that had been treated for 52 weeks and sacrificed at 52 weeks.
    These data indicate that the tumour yield is not dependent upon the
    total dose administered, but rather that sufficient time simply had to
    be provided for the tumours to become evident.

         The carcinogenicity of bromate has also been studied in three
    hybrid strains of mice (B6C3F1, BDF1 and CDF1). Treatments of
    female mice were conducted at concentrations of 0, 500 or
    1000 mg/litre for 78 weeks (average dose 0, 43.5 or 91.6 mg of bromate
    per kg of body weight per day). No treatment-related increases in
    tumour incidence were observed (Kurokawa et al., 1986b). Groups of 27

	FIGURE 8
	FIGURE 9
    

    male mice of the same strains were provided 750 mg of potassium
    bromate per litre (approximately 60-90 mg/kg of body weight per day,
    as cited in FAO/WHO, 1993) for 88 weeks. A control group of 15 males
    per strain was used. Increased numbers of renal cell tumours were not
    observed in any of the strains. There was, however, an increased
    frequency of adenomas (14/27 mice) relative to control mice (1/15) of
    the CDF1 hybrid (Kurokawa et al., 1990).

         In a separate study, male Syrian golden hamsters were treated
    with potassium bromate at concentrations of 0, 125, 250, 500 or 2000
    mg/litre of drinking-water for 89 weeks (Takamura et al., 1985). No
    renal cell tumours were observed in control (0/20) or 125 mg/litre
    (0/19) groups. At the higher concentrations, the incidences were 1/17
    at 250, 4/20 at 500 and 2/19 at 2000 mg/litre.

         DeAngelo et al. (1998) administered potassium bromate to male
    F344 rats and male B6C3F1 mice (78 per group) in drinking-water at
    concentrations of 0, 20, 100, 200 or 400 mg/litre or 0, 80, 400 or
    800 mg/litre, respectively, for 100 weeks. Time-weighted mean daily
    doses were calculated by the authors from mean daily water consumption
    and the measured concentrations of potassium bromate. For rats, six
    animals per group were included for interim sacrifices, which occurred
    at 12, 26, 52 and 77 weeks. Statistically significant, dose-dependent
    increases in tumour incidence were observed in the kidney (adenomas
    and carcinomas combined and carcinomas alone), thyroid (adenomas and
    carcinomas combined and carcinomas alone) and tunical vaginalis testis
    (mesotheliomas). Historical control incidences for these tumour sites
    in male F344 rats are as follows: renal cell tumours, 0.6%; thyroid
    follicular cell adenomas and carcinomas, 2.1%; and mesotheliomas,
    1.5%. The dose-response information for the renal cell tumours and
    mesotheliomas is provided in Figure 7 to facilitate comparisons with
    the Kurokawa studies. The combined thyroid cell tumour incidences were
    0/36 (0%), 4/39 (10%), 1/43 (2%), 4/35 (11%) and 14/30 (47%) in the 0,
    0.1, 6.1, 12.9 and 28.7 mg of bromate per kg of body weight per day
    dose groups, respectively. Thyroid tumours were not reported in the
    Kurokawa studies discussed above.

         As seen in Figure 7, there is remarkable correspondence in the
    dose-response relationships for renal cell tumour induction by bromate
    observed in the DeAngelo et al. (1998) and the Kurokawa et al. (1986a)
    studies. There was no indication of a positive trend in the
    dose-response curve at the lowest doses. The DeAngelo et al. (1998)
    study produced a positive trend in mesothelioma induction at the lower
    dose, but this response was not significant until the dose was raised
    to 6.1 mg/kg of body weight per day, with a strong positive trend in
    tumour incidence with additional doses. This is different from the
    Kurokawa studies, in that a significant background incidence of
    mesotheliomas was observed in one study (Kurokawa et al., 1983),
    whereas there was no indication of spontaneous lesions in a second
    study (Kurokawa et al., 1986a). When the studies are combined, a clear
    positive response is observed at concentrations of 200 mg/litre or
    more in drinking-water (>12.9 mg/kg of body weight per day).

         Tumour responses in male B6C3F1 mice were confined to kidney
    tumours, but the incidence was not clearly dose-dependent. The tumour
    incidence at terminal sacrifice was 0/40 (0%), 5/38 (13%), 3/41 (7%)
    and 1/44 (2%) in mice treated with the equivalent of 0, 6.9, 33 or
    60 mg/kg of body weight per day (DeAngelo et al., 1998).

         A series of additional studies have evaluated the ability of
    bromate to act as a tumour promoter in a variety of animal models.
    Bromate was found to be inactive when applied to the skin of
    dimethylbenzanthracene-initiated female Sencar mice in 0.2 ml of
    acetone at 40 mg of potassium bromate per ml twice per week for 51
    weeks (Kurokawa et al., 1984). Potassium bromate was administered for
    24 weeks at concentrations of 15, 30, 60, 125, 250 or 500 mg/litre to
    male F344 rats initiated by EHEN given in the drinking-water for the
    first 2 weeks at a concentration of 500 mg/litre (Kurokawa et al.,
    1985b). A dose-related increase in renal cell tumours was observed in
    rats treated with more than 30 mg of potassium bromate per litre.
    These changes were apparently not analysed for dose-relatedness, but
    individual groups were compared using the Student  t-test and
    considered statistically insignificant ( p = 0.05). There were,
    however, statistically significant (again compared using the Student
     t-test) increases in the mean number of dysplastic foci at doses of
    30 mg/litre and above ( p < 0.01) and in the mean number of renal
    cell tumours observed per cm2 at 500 mg/litre. These data provide
    evidence that at least some of the carcinogenic activity of bromate
    can be attributed to its activity as a promoter.

    4.8.4  Comparative pharmacokinetics and metabolism

         The absorption and distribution of bromate and bromide were
    studied following administration of a single dose of 50 mg of bromate
    per kg of body weight (Kurokawa et al., 1990). Concentrations of
    bromate of approximately 4 g/ml were seen in plasma 15 min after
    administration. This concentration was quickly reduced to
    approximately 1 g/ml within another 15 min and was not detectable in
    plasma at 2 h. Concentrations of bromate in urine peaked at
    approximately 1 h after administration. No bromate was detected in
    urine until doses reached 5 mg/kg of body weight. About 3-6% of the
    higher doses were recovered in the urine.

         These observations are generally consistent with the observed
    elimination of bromate in human urine following poisonings
    (Lichtenberg et al., 1989). However, the data available in humans are
    quite limited and generally involved much higher doses than would be
    encountered in drinking-water.

    4.8.5  Mode of action

         Attempts have been made to link the toxic effects of bromate with
    its oxidant properties. The administration of potassium bromate
    induces TBARS as a measure of lipid peroxidation (Kurokawa et al.,
    1987b). Single doses of 77 mg/kg of body weight and higher
    significantly increased TBARS in the kidney of F344 rats. Mice
    displayed smaller and less consistent responses, with statistically

    significant responses being seen in male CDF1 mice, but not B6C3F1
    or BDF1 mice. Lipid peroxidation was not observed in male Syrian
    golden hamsters. Treatment of rats with antioxidants (GSH or cysteine)
    decreased the lethality of potassium bromate. Clinical indicators of
    kidney damage (non-protein nitrogen, BUN and creatinine) were
    consistently reduced by GSH or cysteine and increased by
    co-administration of diethylmaleate, a depletor of GSH. These
    treatments modified the histopathological changes in renal tubules
    that were consistent with the clinical findings.

          In vitro studies have shown that hydroxyl radical is produced
    in renal homogenates or kidney cells treated with potassium bromate
    (Sai et al., 1992a,b). Incubation of hepatocytes or liver homogenates
    with bromate did not produce evidence of oxygen radicals.

         Several studies have demonstrated the formation of 8-OH-dG in the
    DNA of the kidneys of rats treated with bromate (Kasai et al., 1987;
    Sai et al., 1991; Cho et al., 1993). Such modifications in DNA can be
    produced by oxygen radicals. The formation of 8-OH-dG was not observed
    in liver DNA of the same animals, paralleling the target organ
    specificity of the compound. Increases in 8-OH-dG were blocked by
    parallel treatment with GSH, cysteine or vitamin C (Sai et al.,
    1992c). Superoxide dismutase and vitamin E were ineffective in
    modifying DNA damage produced by potassium bromate.

         GSH and cysteine administered 30 min before or after bromate
    significantly inhibited the induction of micronuclei in rat peripheral
    blood reticulocytes (Sai et al., 1992c). Treatment of superoxide
    dismutase had no effect on the induction of micronucleated
    reticulocytes. This suggests that the mechanisms involved in
    micronuclei formation parallel those involved in production of lipid
    peroxidation and DNA damage via the production of oxygen radicals.

         Chipman et al. (1998) studied the production of DNA oxidation
    with bromate  in vitro and with intraperitoneal doses of potassium
    bromate. Studies with isolated calf thymus DNA demonstrated a
    GSH-dependent oxidation of guanosine bases.  In vivo studies
    indicated that this mechanism was apparently not active. High doses of
    potassium bromate (100 mg/kg of body weight) administered
    intraperitoneally induced a statistically significant increase in
    8-OH-dG adducts in total cellular DNA. A trend towards an increase was
    observed at 20 mg/kg of body weight per day.

         Studies demonstrating that clastogenic effects of bromate can be
    suppressed by antioxidant treatments provide evidence that effects in
    the bone marrow can be attributed to generation of oxygen radicals and
    the Fenton chemistry that subsequently occurs. The acute doses
    required to produce evidence of damage to DNA mediated by oxygen
    radicals generated by bromate have been considerably higher than the
    daily oral doses required to induce a significant tumour response in
    the rat kidney (6.1 mg/kg of body weight per day). Consequently, more
    precise dose metrics and better dose-response information will be
    necessary to demonstrate the relationship of these effects of bromate
    with tumour induction. The possibility that increased rates of cell

    replication could contribute to the renal cancer induced by bromate
    was investigated by Umemura et al. (1993). Both sodium and potassium
    salts were utilized and were found to induce significant increases in
    the cumulative replication fraction of cells in the proximal
    convoluted tubules of male F344 rats. The effect was significantly
    smaller in proximal straight or distal tubules. However, female rats
    did not display elevated rates of replication (Umemura et al., 1993).
    Thus, these effects in male rats appear to be associated with the
    formation of bodies that are stained with Mallory-Heidenham stain,
    used to detect hyaline droplets. These bodies were not seen in female
    rats. The absence of these responses in female rats, despite a very
    similar incidence of renal tumours in lifetime exposures (Figure 7),
    renders an argument of alpha-2u globulin in the target tissue
    immaterial. While such pathology could contribute to the tumorigenic
    response in the male, it is clearly unnecessary for the response in
    females.

    4.9  Other DBPs

         Many other DBPs can be found in drinking-water, as indicated in
    chapter 2. Most of these are present at very low concentrations.
    Several were considered by WHO (1993, 1996, 1998) in the  Guidelines 
     for drinking-water quality. Those considered in the Guidelines but
    not in this document include formaldehyde, chlorophenols, chloropicrin
    and cyanogen chloride. There are no new data that materially change
    those evaluations.

	

Please Note: This is a corrigendum after the task group

Note: After the printing of the document, Dr James Huff kindly brought to the attention of the Secretariat that a study on the carcinogenicity of sodium hypochlorite, and another on the carcinogenicity of bromodichloromethane, chlorodibromomethane, bromoform, chlorine, and chloramine, were not cited in the document. The authors' abstracts of these studies are given below. Soffritti M, Belpoggi F, Lenzi A, Maltoni C (1997) Results of long-term carcinogenicity studies of chlorine in rats. Ann NY Acad Sci, 837: 189-208. Four groups, each of 50 male and 50 female Sprague-Dawley rats, of the colony used in the Cancer Research Center of Bentivoglio of the Ramazzini Foundation, 12 weeks old at the start of the study, received drinking water containing sodium hypochlorite, resulting in concentrations of active chlorine of 750, 500, and 100 mg/l (treated groups), and tap water (active chlorine < 0.2 mg/l) (control group), respectively, for 104 weeks. Among the female rats of the treated groups, an increased incidence of lymphomas and leukemias has been observed, although this is not clearly dose related. Moreover, sporadic cases of some tumors, the occurrence of which is extremely unusual among the untreated rats of the colony used (historical controls), were detected in chlorine-exposed animals. The results of this study confirm the results of the experiment of the United States National Toxicology Program (1991), which showed an increase of leukemia among female Fischer 344/N rats following the administration of chlorine (in the form of sodium hypochlorite and chloramine) in their drinking water. The data here presented call for further research aimed at quantifying the oncogenic risks related to the chlorination of drinking water, to be used as a basis for consequent public health measures. Dunnick JK, Melnick RL (1993) Assessment of the carcinogenic potential of chlorinated water: experimental studies of chlorine, chloramine, and trihalomethanes. J Natl Cancer Inst, 85: 817-822. BACKGROUND: Water chlorination has been one of the major disease prevention treatments of this century. While epidemiologic studies suggest an association between cancer in humans and consumption of chlorination byproducts in drinking water, these studies have not been adequate to draw definite conclusions about the carcinogenic potential of the individual byproducts PURPOSE: The purpose of this study was to investigate the carcinogenic potential of chlorinated or chloraminated drinking water and of four organic trihalomethane byproducts of chlorination (chloroform, bromodichloromethane, chlorodibromomethane, and bromoform) in rats and mice. METHODS: Bromodichloromethane, chlorodibromomethane, bromoform, chlorine, or chloramine was administered to both sexes of F344/N rats and (C57BL/6 x C3H)F1 mice (hereafter called B6C3F1 mice). Chloroform was given to both sexes of Osborne-Mendel rats and B6C3F1 mice. Chlorine or chloramine was administered daily in the drinking water for 2 years at doses ranging from 0.05 to 0.3 mmol/kg per day. The trihalomethanes were administered by gavage in corn oil at doses ranging from 0.15 to 4.0 mmol/kg per day for 2 years, with the exception of chloroform, which was given for 78 weeks. RESULTS: The trihalomethanes were carcinogenic in the liver, kidney, and/or intestine of rodents. There was equivocal evidence for carcinogenicity in female rats that received chlorinated or chloraminated drinking water; this evidence was based on a marginal increase in the incidence of mononuclear cell leukemia. Rodents were generally exposed to lower doses of chlorine and chloramine than to the trihalomethanes, but the doses in these studies were the maximum that the animals would consume in the drinking water. The highest doses used in the chlorine and chloramine studies were equivalent to a daily gavage dose of bromodichloromethane that induced neoplasms of the large intestine in rats. In contrast to the results with the trihalomethanes, administration of chlorine or chloramine did not cause a clear carcinogenic response in rats or mice after long-term exposure. CONCLUSION: These results suggest that organic byproducts of chlorination are the chemicals of greatest concern in assessment of the carcinogenic potential of chlorinated drinking water. 5. EPIDEMIOLOGICAL STUDIES This chapter reviews observational and experimental epidemiological studies that have been conducted to determine associations between disinfected drinking-water and adverse health outcomes. Disinfection practices vary throughout the world. Applied and residual concentrations have varied over the years and from country to country. Epidemiological study designs, sources of systematic and random error (bias), and guidelines for assessing the causality of associations are discussed in section 5.1. Epidemiological studies of exposures to disinfected drinking-water and to specific DBPs are evaluated in sections 5.2 and 5.3, respectively. Observational epidemiological studies have been conducted to determine possible associations between adverse health-related outcomes and drinking-water disinfected with chlorine and chloramine. Chlorinated drinking-water was studied most often, and studies primarily compared health risks associated with chlorinated drinking-water from surface water sources with those associated with unchlorinated drinking-water from groundwater sources. Also studied were specific DBPs, including chloroform and other THMs. Only one study considered DBPs other than THMs. Two studies considered risks that may be associated with chloraminated water and chlorine dioxide. The mutagenic activity of drinking-water, which may represent exposure to the non-volatile, acid/neutral fraction of chlorinated organic material in water, was also considered. Health effects studied included cancer, cardiovascular disease and adverse reproductive and developmental outcomes. Most of the studies focused on bladder cancer risks. Also studied were risks of colon, rectal and other cancers. 5.1 Epidemiological study designs and causality of epidemiological associations Both observational and experimental epidemiological studies have been conducted to assess the health risks associated with drinking-water disinfection (Table 22). 5.1.1 Experimental studies Results of experimental epidemiological studies, which include clinical trials, are reported in chapter 4 as appropriate under toxicity in humans. These studies consider the effect of varying some characteristic or exposure that is under the investigator's control, much like in a toxicological study. Comparable individuals are randomly assigned to a treatment or intervention group and observed for a specific health-related outcome. Ethical concerns must be fully addressed. Several clinical trials have evaluated changes in lipid, thyroid and haematological parameters that may be affected by consumption of disinfected drinking-water. Table 22. Types of epidemiological studiesa I. Experimental A. Clinical B. Population II. Observational A. Descriptive 1. Disease surveillance and surveys 2. Ecological B. Analytical 1. Longitudinal a. Cohort (follow-up) b. Case-control (case-comparison) 2. Cross-sectional a Adapted from Monson (1990). 5.1.2 Observational studies Two basic kinds of observational epidemiological studies have been conducted to determine risks associated with disinfection of drinking-water: ecological and analytical. These two study approaches differ primarily in the supportive evidence they can provide about a possible causal association. Unlike the analytical study, an ecological study does not link individual outcome events to individual exposure or confounding characteristics, and it does not link individual exposure and confounding characteristics to one another. In an ecological study, information about exposure and disease is available only for groups of people, and critical information can be lost in the process of aggregating these data (Piantadosi, 1994). Results from ecological studies are difficult to interpret, and serious errors can result when it is assumed that inferences from an ecological analysis pertain either to the individuals within the group or to individuals across the groups (Connor & Gillings, 1974; Piantadosi et al., 1988). Theoretical and empirical analyses have offered no consistent guidelines for the interpretation of ecological associations (Greenland & Robins, 1994a,b; Piantadosi, 1994). Investigators (Greenland & Robins, 1994a,b; Piantadosi, 1994; Poole, 1994; Susser, 1994a,b) have examined the limitations of ecological studies and determined when and under what assumptions this type of study may be appropriate. Analytical studies can provide the necessary information to help evaluate the causality of an association and estimate the magnitude of risk. For each person included in the study, information is obtained about their disease status, their exposure to various contaminants and confounding characteristics. Analytical studies are either longitudinal or cross-sectional. In a longitudinal study, the time sequence can be inferred between exposure and disease; in other words, exposure precedes disease. In a cross-sectional study, exposure and disease information relate to the same time period; in these studies, it may not always be correct to presume that exposure preceded disease. The cross-sectional study design was used to investigate possible risks of cardiovascular disease and reproductive and developmental risks. Longitudinal studies are of two opposite approaches: the cohort study and the case-control study. The cohort study begins with the identification of individuals having an exposure of interest and a non-exposed population for comparison; disease consequences or other health-related outcomes are then determined for each group. In a case-control study, the investigator identifies individuals having a disease or health outcome of interest and a control or comparison group of individuals without the disease of interest; exposures and risk factors are evaluated in these persons. In a case-control study, a variety of exposures can be studied, whereas in a cohort study, a variety of diseases can be studied. The cohort or follow-up study can be either retrospective or prospective, and sometimes a combination retrospective-prospective approach is used. Two or more groups of people are assembled for study strictly according to their exposure status. Incidence or mortality rates for the disease of interest are compared between exposed and unexposed groups. Multiple disease end-points can be evaluated, but a disadvantage is that large numbers of people must be studied, especially for environmental exposures. Because of the lengthy latent period for cardiovascular disease and cancer, a long follow-up period is required for a prospective cohort, and this is usually not feasible. The retrospective cohort study design was used to evaluate the possible association of chlorinated drinking-water with cancer and cardiovascular disease risks. In a case-control study, persons with the disease of interest (the cases) and persons without this disease (the controls or comparison group) are sampled from either the general population or a special population (e.g., hospitals or a select group) within a specified geographic area. Exposures among the cases are compared with exposures among non-diseased persons. Multiple exposures can be evaluated, and a relatively small number of study participants is needed to obtain reasonably precise estimates of risk associated with environmental exposures. Retrospective exposures must be considered, and, because of the lengthy latency period for cancer and cardiovascular disease, exposures to water sources and contaminants must be assessed over the previous 20-30 years, or perhaps even a person's lifetime. It may be difficult to assess these exposures accurately. Two types of case-control studies have been conducted to investigate associations between disinfected drinking-water and cancer: * Decedent cases and controls without interviewing next-of-kin or survivors for information about residential histories, risk factors and possible confounding characteristics. * Decedent and incident cases and controls using interviews or other methods to obtain information about possible confounding characteristics and document long-term mobility and changes in residences to allow documentation of lifetime exposure to disinfected water. In several studies, a person's intake of tapwater and historical exposures to chlorinated water, chloroform or other THMs were assessed. 5.1.3 Random and systematic error Biases that occur during the design and conduct of a study can lead to a false or spurious association or a measure of risk that departs systematically from the true value. All reported epidemiological associations require evaluation of random and systematic error so that results can be interpreted properly. Systematic error (bias) affects the validity of a study's observed association; random error affects the precision of the estimated magnitude of the risk. Random error is governed by chance and is influenced by the size of the study. The likelihood that a positive association is due to random error can be assessed by calculating the level of statistical significance ( "P" value) or confidence interval (CI). A small P value or a CI that does not include unity (1.0) suggests that chance may be an unlikely explanation for an observed association, but the association may, nevertheless, be spurious because of systematic error. Statistical significance does not imply causality or biological significance, nor does it mean that random error or chance can ever be completely ruled out as a possible explanation for the observed association. Many epidemiologists believe that strict reliance on statistical significance testing is not appropriate (US EPA, 1994a). Potential sources of systematic error include observation, selection, misclassification and confounding biases. When information on exposure and disease is collected by methods that are not comparable for each participant (e.g., selective recall), an incorrect association will be due to observation bias. When the criteria used to enrol individuals in the study are not comparable, the observed association between exposure and disease will be due to selection bias. A wrong diagnosis of disease or assessment of exposure can result in misclassification bias. This type of bias may be randomly distributed (non-differential misclassification bias), which almost always biases study results towards the direction of not observing an effect (or observing a smaller change in risk than may actually be present), or it may be non-random (differential misclassification bias), which can result in either higher or lower estimates of risk, depending on how the misclassification is distributed. Lynch et al. (1989) examined the effects of misclassification of exposure using empirical data from an interview-based case-control study of bladder cancer in Iowa (USA). Bladder cancer risk estimates were found to be higher when more information was known about the study participants' residential history and their possible exposure to chlorinated water sources. This suggests that misclassification bias in epidemiological studies of chlorinated water may be primarily non-differential, underestimating the risk; however, in study areas where residential mobility is different from that in Iowa, the magnitude of risk may be overestimated rather than underestimated. Confounding bias may convey the appearance of an association; that is, a confounding characteristic rather than the putative cause or exposure may be responsible for all or much of the observed association. Although negative confounding bias may occur, concern is usually with positive effects of confounding bias -- is confounding bias responsible for the observed association? Confounding bias is potentially present in all epidemiological studies and should always be evaluated as a possible explanation for an association. Information on known or suspected confounding characteristics is collected to evaluate and control confounding during the analysis. In the design of case-control studies, matching is a technique that is used to prevent confounding bias. For example, if smoking is thought to be a possible confounding characteristic, an equal number or proportion of smoking cases and controls can be selected for study in order to avoid confounding bias by this exposure. Techniques are also available to assess and control confounding during the data analysis. In an experimental epidemiological study, randomization is possible; that is, each individual in the study has an equal or random chance of being assigned to an exposed or unexposed group. Because of this random assignment of exposure, all characteristics, confounding or not, tend to be distributed equally between the selected study groups of different exposure. Procedures in the study's design and conduct are used to prevent or reduce possible bias. If bias has been identified in a study, the direction of the bias can often be determined, but its effect on the magnitude of the association may not. For example, information may be available to determine whether the bias was responsible for an increased or decreased likelihood of observing an association, but its magnitude usually cannot be estimated. Two basic measures of an association between exposure and disease in analytical studies are the rate ratio or relative risk (RR) and exposure odds ratio (OR). A mortality odds ratio (MOR) is reported when mortality is studied. An RR or OR of 1.0 indicates no association; any other ratio signifies either a positive or negative association. For example, an RR or OR of 1.8 indicates an 80% increased risk among the exposed. Decreased risk and protective effects are indicated by an RR or OR that is less than 1.0. The size of the relative risk and odds ratio is also used to help assess if an observed association may be spurious (Table 23). Based on Monson's (1990) experience, an RR or OR of 0.9-1.2 indicates essentially no association. Associations in this range are generally considered too weak to be detected by epidemiological methods. It is difficult to interpret a weak association, i.e., an RR or OR of 1.2-1.5. One or more confounding characteristics can easily lead to a weak association between exposure and disease, and it is usually not possible to identify, measure or control weak confounding bias. On the other hand, a large relative risk is unlikely to be completely explained by some uncontrolled or unidentified confounding characteristic. When the study has a reasonably large number of participants and the relative risk or odds ratio is large, random variability and confounding bias are much less likely to be responsible for an observed association. Table 23. Guide to the strength of an epidemiological associationa Relative risk Strength of association 1.0 None >1.0-<1.5 Weak 1.5-3.0 Moderate 3.1-10.0 Strong >10.0 Infinite a Adapted from Monson (1990). Another measure of effect is the standardized mortality ratio (SMR). An SMR of 100 indicates no association; an SMR of 150 indicates a 50% increased risk. 5.1.4 Causality of an epidemiological association Epidemiological associations may be causal; however, before causality can be assessed, each study must be evaluated to determine whether its design is appropriate, the study size is adequate and systematic bias has not influenced the observed association. In addition, the association should be consistent with prior hypotheses and previous study results, and its magnitude should be moderately large. Causality requires sufficient evidence from several well designed and well conducted epidemiological studies in various geographic areas. Supporting toxicological and pharmacological data are also important. Guidelines are available to help epidemiologists assess the possible causality of associations observed in well designed and well conducted studies. Epidemiological data should be interpreted with caution and in the context of other available scientific information. Epidemiologists apply the following guidelines to assess evidence about causality (Hill, 1965; Rothman, 1986): * Biological plausibility. When the association is supported by evidence from clinical research or toxicology about biological behaviour or mechanisms, an inference of causality is strengthened. * Temporal association. Exposure must precede the disease, and in most epidemiological studies this can be inferred. When exposure and disease are measured simultaneously, it is possible that exposure has been modified by the presence of disease. * Study precision and validity. Individual studies that provide evidence of an association are well designed with an adequate number of study participants (good precision) and well conducted with valid results (i.e., the association is not likely due to systematic bias). * Strength of association. The larger the relative risk or odds ratio, the less likely the association is to be spurious or due to unidentified confounding. However, a causal association cannot be ruled out simply because a weak association is observed. * Consistency. Repeated observation of an association under different study conditions supports an inference of causality, but the absence of consistency does not rule out causality. * Specificity. A putative cause or exposure leads to a specific effect. The presence of specificity argues for causality, but its absence does not rule it out. * Dose-response relationship. A causal interpretation is more plausible when an epidemiological gradient is found (e.g., higher risk is associated with larger exposures). * Reversibility or preventability. An observed association leads to some preventive action, and removal of the possible cause leads to a reduction of disease or risk of disease. 5.2 Epidemiological associations between disinfectant use and adverse health outcomes Studies of water disinfected with chlorine and chloramine are reviewed in this section. Chlorinated drinking-water was studied most often. Studies primarily compared health risks associated with ingestion of chlorinated drinking-water from surface water sources with those associated with ingestion of unchlorinated drinking-water from groundwater sources, but risks were also compared among populations using chloraminated and chlorinated surface water supplies. Studies that considered exposures to specific DBPs are reviewed in section 5.3. Studies that considered exposures to both disinfected drinking-water and specific by-products are described in either section 5.2 or 5.3. One study assessed the effects on haematological and serum chemical parameters that may be associated with the use of chlorine dioxide. Because of their relevance to other experimental studies, results are reported in section 4.6.3. 5.2.1 Epidemiological studies of cancer and disinfected drinking-water Since 1974, numerous epidemiological studies have attempted to assess the association between cancer and the long-term consumption of disinfected drinking-water. Studies were conducted in various geographic locations with different types of water sources, chemical quality and levels and types of DBPs. Ecological, cohort and case-control studies of incident and decedent cases were conducted. The quality of information about water disinfection exposures and potential confounding characteristics differs dramatically between these studies. In many of the case-control studies, interviews or other methods were used to obtain information about various risk factors, confounding characteristics and residential histories, to determine long-term exposures to disinfected drinking-water supplies; in several studies, individual tapwater consumption was estimated. However, in several other case-control studies, limited information about exposure and confounding factors was obtained only from the death certificates. In most studies, disease incidence or mortality was compared between populations supplied with chlorinated surface water and those supplied with unchlorinated groundwater. The chemical quality of drinking-water for a number of chemical constituents, including DBPs, differs between surface water and groundwater and also among the various surface waters in the different geographic locations studied. Surface water sources may also be contaminated with non-volatile synthetic organic compounds from industrial, agricultural and residential runoff. Groundwater may be contaminated with volatile synthetic organic compounds and inorganic constituents, such as arsenic and nitrate. It is not feasible to consider epidemiological studies of cancer in populations consuming unchlorinated drinking-water from surface water sources because so few people consume such drinking-water, but exposures to water contaminants in addition to DBPs must be considered. Since the quality of water sources may also affect the concentration and type of DBPs, even when the same disinfectant is used, it is important to assess specific by-products in water systems included in epidemiological studies. 5.2.1.1 Cancer associations in ecological studies Early epidemiological studies analysed group or aggregate data available on drinking-water exposures and cancer. Usually the variables selected for analysis were readily available in published census, vital statistics or public records. Cancer mortality rates, usually obtained for census tracts, counties or other geographic regions, were compared for areas with different water sources and disinfection practices. Differing source waters and their disinfection were usually used as surrogates for drinking-water exposures of the populations. For example, the drinking-water for the area was categorized as being primarily from a surface water or from a groundwater source and as chlorinated or not chlorinated. In some instances, exposure variables included estimates of the proportion of the area's population that received drinking-water from chlorinated surface water or unchlorinated groundwater. In several ecological analyses, the investigators studied the association between cancer mortality and an estimate of population exposures to levels of chlorination by-products based on the measure of THMs or chloroform levels from a limited number of water samples from monitoring studies (see section 5.3.1.1). The first ecological studies reported higher mortality from several cancers in Louisiana (USA) parishes where the majority of the population used the lower Mississippi River as a source of drinking-water. Lower mortality was reported in parishes where the majority of the population used groundwater sources (Harris, 1974; Page et al., 1976). Additional ecological studies for different geographic areas of the USA, including Louisiana, Ohio, Missouri, Kentucky, New York, Massachusetts and Iowa, reported increased cancer mortality or incidence in areas using chlorinated surface water (Craun, 1991). A wide range of cancer sites was found to be statistically associated with the use of chlorinated surface water. These cancer sites included gall bladder, oesophagus, kidney, breast, liver, pancreas, prostate, stomach, bladder, colon and rectum. These studies have been extensively reviewed (NAS, 1980, 1986; Crump & Guess, 1982; Shy, 1985; Craun, 1991; Murphy & Craun, 1990). The NAS (1980) noted the limitations of these studies and recommended that analytical studies be conducted to further assess the possible association of chlorinated drinking-water and cancer. It was recommended that studies focus on cancers of the bladder, stomach, colon and rectum because they had been most often associated with chlorinated water in ecological studies. As noted in section 5.1.2, ecological analyses have theoretical deficiencies, and the interpretation of results from these studies is difficult. Associations reported from ecological analyses cannot be evaluated for causality, nor do they provide an estimate of the magnitude of risk. Yet these studies continue to be conducted. An ecological study in Norway reported weak associations between chlorinated drinking-water supplies and cancer of the colon and rectum in men and women (Flaten, 1992). No associations between chlorinated surface water supplies and bladder or stomach cancer were found in Valencia province, Spain (Suarez-Varela et al., 1994). A study in New Jersey found no association between DBPs and either bladder or rectal cancer (Savin & Cohn, 1996). A study in Taiwan, China, reported associations between the use of chlorinated drinking-water and cancer of the rectum, lung, bladder and kidney (Yang et al., 1998). THMs were not found to be associated with breast cancer risk in an ecological study in North Carolina (Marcus et al., 1998). 5.2.1.2 Cancer associations in analytical studies More case-control studies have been conducted than cohort studies. Case-control studies have included the traditional interview-based study and those where information was obtained only from death certificates or other readily available sources. 1) Cohort studies Wilkins & Comstock (1981) found no statistically significant associations between the incidence of cancer mortality in Washington County, Maryland, USA, and residence in an area supplied with chlorinated surface water. Information was available for individuals in a well-defined homogeneous cohort that allowed disease rates to be computed by presumed degree of exposure to chlorination by-products. The cohort was established from a private census during the summer of 1963 and followed for 12 years. The source of drinking-water at home was ascertained, and personal and socioeconomic data were collected for each county resident, including age, education, smoking history and number of years lived at the 1963 address. Potential cases of cancer were obtained from death certificate records, the county's cancer registry and medical records of the county hospital and a regional medical centre. Census data were used to compute age-gender-site-specific cancer mortality incidence rates for 27 causes of death, including 16 cancer sites, cardiovascular disease, vehicular accidents, all causes of death and pneumonia. Three exposure categories were examined: a high-exposure group of residents served by chlorinated surface water, a low-exposure group served by unchlorinated deep wells and a third group served by a combination of chlorinated surface water and groundwater. The average chloroform level from an extensive analysis of chlorinated surface water samples was 107 g/litre. The third group, which represented an intermediate exposure, was not used in detailed analyses. Confounding bias was controlled and incidence rates were adjusted by multiple regression analysis for age, marital status, education, smoking history, frequency of church attendance, adequacy of housing and number of persons per room. Selected cancer mortality rates for males and females are reported in Table 24. Small increased risks of bladder and liver cancer were reported; these risk estimates are not statistically stable. Although the study was of high quality and well conducted, the associations reported are subject to random error (i.e., all relative risks had a confidence interval that included 1.0 and thus were not statistically significant). Even though over 31 000 people were included in the cohort, estimates of specific cancer risks were based on relatively few deaths. History of length of residence was used to estimate a person's duration of exposure to chlorinated and unchlorinated water. For bladder and liver cancer in females and bladder cancer in males, the association was stronger for persons who had lived in their 1963 domicile for 12 or more years than for those who had lived in it 3 years or less. Among men who had at least 24 years' exposure to Table 24. Cohort study: Selected cancer mortality, water source and disinfection, Washington County, Maryland (USA)a Cause of death Chlorinated surface water Unchlorinated groundwater Risk Deaths Incidence rateb Deaths Incidence rateb RR 95% CI Females Liver cancer 31 19.9 2 11.0 1.8 0.6-6.8 Kidney cancer 11 7.2 2 7.1 1.0 0.3-6.0 Bladder cancer 27 16.6 2 10.4 1.6 0.5-6.3 Males Liver cancer 9 6.4 2 9.0 0.7 0.2-3.5 Kidney cancer 15 10.6 3 13.6 0.8 0.3-2.7 Bladder cancer 46 34.6 5 19.2 1.8 0.8-4.8 a From Wilkins & Comstock (1981). b Adjusted incidence rate per 100 000 person-years. chlorinated surface water, the bladder cancer risk was high (RR = 6.5; 95% CI = 1.0->100) but very imprecise and statistically unstable (the CI is high). Additional follow-up of the cohort for several more years can possibly provide more meaningful associations. Freedman et al. (1997) and Ijsselmuiden et al. (1992) conducted case-control studies in Washington County, Maryland (USA) to evaluate risks of cancer of the bladder and pancreas. The results of these two studies are given below. Zierler et al. (1986) examined mortality patterns of Massachusetts (USA) residents at least 45 years of age who died between 1969 and 1983 and whose last residence was in a community where drinking-water was disinfected with either chlorine or chloramine. A standardized mortality study found little differences in the patterns of 51 645 deaths due to cancer in 43 Massachusetts communities with water supplies disinfected with either chlorine or chloramine as compared with all cancer deaths reported in Massachusetts. The mortality rate of residents from these selected communities with chlorinated drinking-water was slightly higher than expected for stomach cancer (SMR = 109; 95% CI = 104-114) and lung cancer (SMR = 105; 95% CI = 103-107). Mortality rates in selected communities served by chloraminated drinking-water was slightly less than expected for bladder cancer (SMR = 93; 95% CI = 88-98) but greater than expected for lung cancer (SMR = 104; 95% CI = 102-106). Because residence at death was used to assign the exposure status of persons to either chlorinated or chloraminated drinking-water, there is a serious potential for exposure misclassification bias. In addition, errors in death certificate classification of the cause of death may also affect the interpretation of these findings. Bean et al. (1982) conducted a cohort study of municipalities in Iowa (USA) that were classified into groups based on the source of their drinking-water -- surface water or groundwater of various depths. Municipalities included were those with a 1970 population of more than 999 and a public water supply that used exclusively either surface water or groundwater sources that had remained stable for at least 14 years. All of the surface water sources were chlorinated; groundwater was less frequently chlorinated, especially as the depth of the well increased. In the regression analysis, each group was considered a single population, and age-adjusted, sex-specific cancer incidence rates were determined for the years 1969-1978. Indicators of socioeconomic status were obtained for the municipalities to determine if they could explain any observed differences in cancer incidence, and detailed information about residential mobility, water usage and smoking, collected from a sample of non-cancer controls in a large bladder cancer case-control study, was used to assess exposure misclassification and confounding. Municipalities supplied by chlorinated surface water had higher lung and rectal cancer incidence rates than those using groundwater sources for all population groups (<10 000; 10 000-50 000; >50 000); no differences were found for colon and bladder cancer incidence between surface water and groundwater sources. Using information reported by Bean et al. (1982), Poole (1997) estimated a single relative risk for lung cancer (RR = 1.1; 95% CI = 1.1-1.2) and rectal cancer (RR = 1.2; 95% CI = 1.1-1.3). Selected results from a prospective cohort study of 41 836 post-menopausal women in Iowa are reported in Table 25. The study compared risks among users of groundwater and surface water sources; an assessment of exposure to chloroform and other THMs was also included (Doyle et al., 1997). In 1989, the participants were asked about their source of drinking-water and the length of time this source had been used. Information on potential confounding characteristics was obtained from the baseline questionnaire. Analyses were limited to those who reported drinking municipal or private well-water for more than the past 10 years ( n = 28 237). Historical water treatment and water quality data were used to ascertain exposure to THMs (see section 5.3.1.2). No increased risks were found for women who used private wells. In comparison with women who used 100% municipal groundwater sources, women who used 100% municipal surface water sources were at a statistically increased risk of all cancers combined (RR = 1.3; 95% CI = 1.1-1.6), colon cancer and breast cancer. No increased risk was observed for bladder cancer or cancer of the rectum and anus; however, only two and six cases of cancer were observed in the cohort. No increased risks were found for cancer of the kidney and six other sites. Reported relative risks were adjusted for age, education, smoking status, physical activity, fruit and vegetable intake, total energy intake, body mass index and waist to hip ratio. Limitations of this study include the relatively short period (8 years) of follow-up of the cohort and possible misclassification of exposure because the source of drinking-water was assessed at only one point in time. 2) Decedent case-control studies without interviews or residential histories Six case-control studies where no information was obtained from next-of-kin interviews (Alavanja et al., 1978; Struba, 1979; Brenniman et al., 1980; Gottlieb et al., 1981, 1982; Young et al., 1981; Zierler et al., 1986) considered only information that was routinely recorded on death certificates or readily available from vital statistics, such as occupation, race, age and gender. Deceased cases of cancer of interest were identified, and controls were non-cancer deaths from the same geographic area. In four of these studies, controls were matched for certain characteristics, including age, race, gender and year of death, to prevent possible confounding bias by these characteristics. In studies where matching was not employed, these characteristics were controlled in the analysis. Other possible confounders considered included occupation listed on the death certificate and a measure of urbanization of the person's residence. Information about important potential confounders, such as diet and smoking, was not available and could not be assessed. In one study, smoking was indirectly controlled by assessing lung cancer risks among those exposed to chlorinated water. Table 25. Cohort study: Selected incidence of cancer in post-menopausal women and water source in Iowa (USA)a Cancer 100% municipal Mixed municipal surface water 100% municipal surface water groundwater and groundwater Cases RRb Cases RRb 95% CI Cases RRb 95% CI Bladder 23 1.0 16 2.4 1.3-4.6 2 0.7 0.2-2.9 Colon 106 1.0 47 1.5 1.1-2.2 23 1.7 1.1-2.7 Rectal 53 1.0 20 1.3 0.8-2.2 6 0.9 0.4-2.1 Breast 381 1.0 106 0.9 0.8-1.2 65 1.3 1.03-1.8 Kidney 21 1.0 5 0.8 0.3-2.2 3 1.1 0.3-3.8 Lung 95 1.0 30 1.1 0.7-1.1 17 1.4 0.9-2.4 Melanoma 29 1.0 12 1.4 1.7-2.8 4 1.1 0.4-3.2 All cancers 631 1.0 223 1.2 1.04-1.4 112 1.3 1.1-1.6 a From Doyle et al. (1997). b Reported relative risks were adjusted for age, education, smoking status, physical activity, fruit and vegetable intake, total energy intake, body mass index and waist to hip ratio. A single address (address at death or usual address) or combination of addresses (address at birth and death) was used to determine place of residence for assessing exposure to disinfected water. This place of residence was linked to public records of water source and treatment practices to classify the type of water source and drinking-water disinfection. Exposures considered were surface water or groundwater sources and chlorinated or unchlorinated water sources at the residence (Shy, 1985); in one study, chlorinated and chloraminated water were evaluated (Zierler et al., 1986). In three studies, a statistically significant increased risk of bladder cancer was found to be associated with chlorinated surface water; an increased risk of colon cancer was found in three studies, and an increased risk of rectal cancer was found in four studies (Shy, 1985; Zierler et al., 1986). Alavanja et al. (1978) studied 3446 deaths due to total urinary tract and total gastrointestinal cancers in seven New York (USA) counties during 1968-1970. The comparison group was non-cancer deaths from the same period matched on age, race, gender, county of birth and county of residence; occupation and the urban nature of the county were considered as potential confounders. Statistically significant increased risks of colon (OR = 2.0) and bladder (OR = 2.0) cancer mortality were found to be higher in men but not women who resided in communities with chlorinated surface water. Other statistically significant increased risks in men who resided in communities with chlorinated surface water were for liver and kidney (OR = 2.8), oesophagus (OR = 2.4) and pancreas (OR = 2.2) cancer mortality; stomach cancer risks were increased for both men (OR = 2.4) and women (OR = 2.2). A study by Struba (1979) of bladder, colon and rectal cancer deaths (700-1500 cases per site) and a comparison group of deaths matched on age, race, gender and region of residence in North Carolina (USA) during 1975-1978 also found statistically significant increased risks for bladder (OR = 1.5), rectal (OR = 1.5) and colon (OR = 1.3) cancer mortality associated with chlorinated water. In studies of similar design, Brenniman et al. (1980), Young et al. (1981) and Gottlieb et al. (1981, 1982) did not find significant increased risks of bladder or colon cancer mortality associated with chlorinated water. Gottlieb et al. (1981, 1982) found increased risks for rectal (OR = 1.7) and breast (OR = 1.6) cancer mortality associated with chlorinated water in Louisiana (USA). In Illinois (USA), Brenniman et al. (1980) found an increased risk of rectal cancer mortality, but only for females (OR = 1.4). Young et al. (1981) reported an association between colon cancer mortality in Wisconsin (USA) and the average daily chlorine dosage of drinking-water over a 20-year period. The study included 8029 cancer deaths and 8029 non-cancer deaths in white women matched on county of residence, age and year of death. Urbanization, marital status and occupation were considered as potential confounders. Logistic regression analysis was used to evaluate risks for a number of site-specific cancer deaths associated with drinking-water classified by investigators as high, medium and low chlorine-dosed water. Only cancer of the colon was found to be statistically associated with the use of chlorine, but the risk did not increase with higher chlorine dosages. Young et al. (1981) found no association between chlorinated drinking-water and mortality from cancer of the bladder, liver, kidney, oesophagus, stomach, pancreas, lung, brain or breast. Zierler et al. (1986) evaluated exposures to surface water supplies disinfected with either chlorine or chloramine among 51 645 persons aged 45 years and older who died from cancer and 214 988 controls who died from cardiovascular, cerebrovascular or pulmonary disease or from lymphatic cancer in 43 Massachusetts (USA) communities. Using lymphatic cancer deaths as the comparison group, bladder cancer mortality was found to be moderately increased (OR = 1.7; 95% CI = 1.3-2.2) among residents who died in communities with chlorinated drinking-water as compared with those who died in communities with chloraminated water. This analysis controlled for the potential effects of differences in population density, poverty, age at death and year of death between the communities treated with different disinfectants. Because smoking is a known risk factor for bladder cancer, a special comparison group of lung cancer deaths was enrolled to evaluate this potential cause of confounding bias. Results suggested that smoking did not explain all of the excess mortality from bladder cancer; thus, chlorinated drinking-water may present some risk. However, because residence at death is a poor measure of long-term exposure to disinfected water, an interview-based case-control study was designed to further evaluate the possible association between chlorinated water and bladder cancer (see below). Although not subject to all of the design limitations of ecological studies, these decedent case-control studies are, nevertheless, still limited in their ability to provide information about the causal nature of the associations observed. Interpretation of results reported by these studies is limited because of likely systematic bias due to misclassification and uncontrolled confounding. Use of decedent instead of incident cancer cases means that differential survival among cases may also influence the observed association. Selection bias is also likely in the control group. Insufficient information was available in these studies to adequately assess historical, long-term exposures to chlorinated water, and use of a single residential address can result in exposure misclassification bias, which may lead to risk being underestimated or overestimated. While the magnitude of risk may be underestimated, as was the case in Iowa (Lynch et al., 1989), risk may also be overestimated as a result of possible different residential mobility patterns. Thus, these studies can provide very limited information about causality and the magnitude of cancer risks of chlorinated water. The findings from these studies should be interpreted with caution because of study design limitations. 3) Case-control studies with interviews or residential histories Two additional decedent case-control studies included more complete information about residential histories for a better assessment of exposure to disinfected water. Lawrence et al. (1984) studied the relationship of THMs and colo-rectal cancer among a cohort of white female teachers in New York State (USA) who died of either colo-rectal cancer or non-cancer causes. A case-control study (Zierler et al., 1990) of individuals who had died of primary bladder cancer and other causes was conducted in selected Massachusetts (USA) communities that obtained drinking-water from surface water sources disinfected by either chlorine or chloramine. Eleven case-control studies assessed risks of cancer incidence and included interviews with study participants or a surrogate to obtain information about complete residential histories, long-term drinking-water exposures to chlorinated or chloraminated water and potential confounding characteristics (Cantor et al., 1985, 1987, 1990, 1995, 1997, 1998; Cragle et al., 1985; Young et al., 1987, 1990; Lynch et al., 1990; Ijsselmuiden et al., 1992; McGeehin et al., 1993; Vena et al., 1993; King & Marrett, 1996; Freedman et al., 1997; Hildesheim et al., 1998). In seven of these studies (Cantor et al., 1987, 1990, 1995, 1997, 1998; McGeehin et al., 1993; Vena et al., 1993; King & Marrett, 1996; Hildesheim et al., 1998), information about tapwater consumption was also obtained. In six studies (Young et al., 1987, 1990; McGeehin et al., 1993; Cantor et al., 1995, 1997, 1998; Hildesheim et al., 1998; King & Marrett, 1996), THM exposure was evaluated (see also section 5.3.1.2). Bladder cancer risk Chlorinated water studies. Cantor et al. (1985) reported results from a national study of water chlorination and bladder cancer in the USA, the largest case-control study of water chlorination risks reported to date. The study included 2982 people between the ages of 21 and 84 diagnosed with bladder cancer in 1978 and residing in five states and five metropolitan areas of the USA and 5782 population-based comparison subjects, randomly selected and frequency matched on gender, age and study area. The primary purpose of the study was to evaluate the possible association of bladder cancer and artificial sweeteners; however, because of its case-control design, it was possible to include an assessment of drinking-water exposures. Participants were interviewed for information about past residences, smoking, occupation, artificial sweetener use, coffee and tea consumption and use of hair dyes. A lifetime residence history categorized individuals according to water sources and chlorination status on a year-by-year basis. Information was obtained on use of bottled water and fluid consumption. Logistic regression analysis was used to control for potential confounding bias. Overall, no association was found between bladder cancer risk and duration of exposure to chlorinated water. In all study areas combined, no increased risk was found among participants who lived in areas with chlorinated water supplies for <20, 20-39, 40-59 and 60 or more years (Table 26A). However, an increased bladder cancer risk was found among persons who never smoked, were never employed in a high-risk occupation and resided in areas served by chlorinated surface sources (Table 27A). There was little evidence of an exposure-response relationship among persons who had never smoked. Only in the low-risk group of non-smokers who had resided 60 or more years in an area served by chlorinated water was the increased bladder cancer risk statistically significant (RR = 2.3; 95% CI = 1.3-4.2). Although the study included a large number of cases and controls, this subgroup analysis included relatively few study participants (Table 27A). Only 46 cases and 77 controls were included in the analysis, which found a doubling of risk among non-smokers who had resided 60 or more years in an area served by chlorinated water. Lynch et al. (1990) conducted a separate analysis of the Iowa (USA) portion of the national bladder cancer study. Included were 294 primary, histologically confirmed cases of bladder cancer in whites and 686 comparison subjects, all of whom had spent more than 50% of their lifetimes on primary water sources with known chlorination exposure. Study participants exposed to chlorinated water sources for more than 40 years were found to have twice the risk of bladder cancer (OR = 2.0; 95% CI = 1.3-3.1) compared with participants exposed to unchlorinated groundwater sources. Unadjusted risks were higher (OR = 9.9; 95% CI = 2.6-38.0) when the analysis was restricted to study participants who used, for more than 40 years, water sources that had been chlorinated prior to filtration, a water treatment practice known to result in higher THM levels. However, for both of these chlorination practices, no statistically significant risks were found to be associated with the use of chlorinated water for 40 or fewer years. The risk of bladder cancer was also increased for cigarette smokers with longer duration of exposure to chlorinated drinking-water. Unadjusted risks were 4 times as great in heavy smokers (OR = 9.9; 95% CI = 2.6-38.0) as in non-smokers (OR = 2.7; 95% CI = 1.2-5.8) who were exposed to chlorinated water for more than 30 years. A greater risk of bladder cancer was also seen in heavy smokers (OR = 4.0; 95% CI = 2.0-8.4) who were exposed to chlorinated water for 30 or fewer years. King & Marrett (1996) conducted a population-based case-control study in Ontario (Canada). Cases were residents between 25 and 74 years of age with a histologically confirmed diagnosis of primary cancer or carcinoma in situ of the bladder, diagnosed between 1 September 1992 and 1 May 1994. Of 1694 eligible cases, 250 patients were not included in the study because an appropriate physician could not be identified or the patients were recently deceased or too ill to be contacted. Of the remaining 1444 patients, consent was received to contact 1262 (87%). The remaining 13% of cases did not participate because the physician refused to participate, did not provide consent or did not respond. Controls were an age-sex frequency matched sample of the general population from households randomly selected from a computerized database of residential telephone listings in the same area. Since the control subjects were also used to study cancers of the colon and rectum with respect to the same exposures (results not yet published), they were selected to have the expected age-sex Table 26. Bladder cancer risks and duration of exposure to chlorinated surface water in five interview-based incident case-control studies Years of Odds 95% CI Comments Reference exposure to ratio chlorinated surface water A. Ten areas of the USA Whites only; adjusted Cantor et al. (1985) 0 1.0 for age, gender, smoking, 1-19 1.1 0.8-1.4 usual employment as a 20-39 1.0 0.8-1.3 farmer, study area 40-59 1.0 0.8-1.3 >60 1.1 0.8-1.5 B. Ontario, Canada Adjusted for age, gender, King & Marrett (1996) 0-9 1.0 smoking, education, calorie 10-19 1.0 0.7-1.5 intake 20-34 1.2 0.9-1.5 >35 1.4 1.1-1.8 C. Colorado Whites only; adjusted for McGeehin et al. (1993) 0 1.0 gender, smoking, coffee 1-10 0.7 0.4-1.3 consumption, tapwater 11-20 1.4 0.8-2.5 intake, family history 21-30 1.5 0.8-2.9 of bladder cancer, medical >30 1.8 1.1-2.9 history of bladder infection or kidney stone D. Iowa Adjusted for age, gender, Cantor et al. (1998) 0 1.0 smoking, education, high-risk <19 1.0 0.8-1.2 employment, study area 20-39 1.1 0.8-1.4 40-59 1.2 0.8-1.7 >60 1.5 0.9-2.6 E. Washington County, Maryland Adjusted for age, gender, Freedman et al. (1997) 0 1.0 smoking, urbanicity 1-10 1.0 0.6-1.5 11-20 1.0 0.6-1.6 21-30 1.1 0.6-1.8 31-40 1.1 0.6-2.2 >40 1.4 0.7-2.9 Table 27. Bladder cancer risks for smokers and non-smokers exposed to chlorinated water Years at a residence Cases Controls Odds ratio Cases Controls Odds ratio Reference/comments served by chlorinated (95% CI) (95% CI) water A. Males and females, Never Current Cantor et al. (1985) 10 areas of the USA smoked smoker 0 61 268 1.0 87 109 1.0 Whites, adjusted for study 1-19 29 110 1.3 (0.7-2.2) 63 71 0.9 (0.6-1.5) area, gender, age, usual 20-39 73 236 1.5 (0.9-2.4) 136 186 0.7 (0.4-1.1) employment as a farmer, 40-59 108 348 1.4 (0.9-2.3) 166 211 0.7 (0.5-1.2) smoking >60 46 77 2.3 (1.3-4.2) 27 37 0.6 (0.3-1.2) B. Males and females, Non-smokers Smokers McGeehin et al. (1993) Colorado 0 19 45 1 85 57 1 Whites only; adjusted for 1-11 7 21 0.8 (0.2-2.4) 36 33 0.8 (0.4-1.4) gender, smoking, coffee 12-34 11 34 0.9 (0.3-2.3) 70 34 1.4 (0.8-2.4) consumption, tapwater intake, >34 21 25 2.9 (1.2-7.4) 77 25 2.1 (1.1-3.8) family history of bladder cancer, medical history of bladder infection or kidney stone C. Males and females, Never Smoker (Freedman et al., 1997) Washington County, smoked (past/present) Maryland 0 32 331 1.0 47 391 1.3 (0.8-2.2) Adjusted for age, gender, and 1-10 21 232 0.8 (0.4-1.6) 70 469 1.4 (0.8-2.5) urbanicity 11-20 15 147 0.9 (0.5-1.9) 41 285 1.4 (0.8-2.5) 21-30 6 89 0.6 (0.2-1.5) 32 177 1.7 (0.9-3.2) 31-40 3 53 0.5 (0.1-1.5) 13 54 2.2 (1.0-4.7) >40 5 51 0.9 (0.3-2.3) 8 27 2.8 (1.0-6.9) Table 27. (continued) Years at a residence Cases Controls Odds ratio Cases Controls Odds ratio Reference/comments served by chlorinated (95% CI) (95% CI) water D. Males, Iowa Never Current Cantor et al. (1998) smoked smoker 0 112 332 1.0 188 156 3.5 (2.5-4.7) Adjusted for age, study period, 1-19 27 75 1.0 (0.6-1.6) 73 62 3.5 (2.3-5.3) education, high-risk population 20-39 6 22 0.8 (0.3-2.0) 37 20 5.7 (3.1-10.4) >40 5 26 0.7 (0.3-1.9) 29 16 5.8 (3.0-11.3) distribution of these three cancer sites combined. In over 90% of the 2768 households with an eligible resident, the person selected for the study agreed to participate. The overall response rate considering actual participation was 73% for cases and 72% for controls. Relevant exposure and confounding variables were collected using a mailed questionnaire in combination with a computer-assisted telephone interview. Interviewers were blinded as to the case or control status of the subject. Questions were included on demographics (e.g., gender, date of birth and education), other potentially important confounding variables (e.g., smoking history and usual diet prior to diagnosis) and information pertaining to the primary exposures of interest (e.g., residence and water source history and usual water consumption prior to diagnosis). Persons reported their drinking-water source and other household water sources at each residence as municipal, household well, bottled water or other. Volume of tapwater was calculated from the reported daily frequency of consuming beverages containing water during the 2-year period before the interview and usual source of water (tap or bottled) used to make hot and cold beverages. The water supply for each participant was characterized by source (surface water vs. groundwater) and chlorination status (chlorinated vs. unchlorinated). The analysis considered the 696 cases (75%) and 1545 controls (73%) for whom water source characteristics were available for at least 30 of the 40 years ending 2 years before the person's interview. To reduce possible misclassification of exposure, only persons with 30 or more years of known water history were included in a separate analysis. Logistic regression was used to estimate bladder cancer risks. Potentially important confounders considered were age, gender, smoking, education, consumption of alcoholic beverages, coffee consumption, total fluid consumption and dietary intake of energy (total calories), protein, fat, cholesterol, fibre and vitamin A. The study found a pattern of increasing bladder cancer risk with increasing number of years exposed to chlorinated surface water, but a statistically significant association was found only for lengthy exposures (Table 26B). For persons exposed to a chlorinated surface water source for 35 years or more, the bladder cancer relative risk was increased by 40% in comparison with those exposed for less than 10 years. An analysis restricted to those who had relatively homogenous water exposures found that exposure to chlorinated surface water for 30 or more years was associated with a weak increased risk (OR = 1.4; 95% CI = 1.1-1.8) compared with exposure to a groundwater source. King & Marrett (1996) found higher relative risk estimates for non-smokers associated with many years of exposure to chlorinated drinking-water from surface water sources, but the difference in risk compared with smokers was not statistically significant, nor was the pattern of higher risks in non-smokers observed consistently. McGeehin et al. (1993) conducted a case-control study to assess the relationship between chlorinated or chloraminated drinking-water and bladder cancer in Colorado. Study participants were identified from a population-based cancer registry; 327 histologically confirmed incident bladder cancer cases and 261 randomly selected controls with other cancers, except colo-rectal and lung, were interviewed by a single blinded interviewer about demographic data, drinking-water source, fluid consumption and personal habits. Persons with colo-rectal and lung cancers were excluded from the control group because of a possible association with chlorinated surface water reported in earlier studies. Of the originally identified cases, 38% could not be interviewed because permission was denied by the physician, cases were dead or persons refused to participate. For controls, 47% were not interviewed for these reasons. Because of the large number of excluded study participants, selection bias is a concern. Bias could be present if the excluded persons were different from participating persons in ways that are related to the exposure and outcome. Information was collected from individuals on tapwater consumption and from local water utilities regarding water source and method of disinfection in order to construct individual lifetime profiles of exposure to disinfected drinking-water. Although 91% of person-years of community exposure to specific water sources could be determined, only 81% of disinfection methods, 60% of THM levels and 55% of residual chlorine levels could be determined for use in assessing exposures. Logistic regression analysis, adjusting for coffee consumption, smoking, tapwater intake, family history of bladder cancer, sex and medical history of bladder infection or kidney stone, found no statistically significant increased risk of bladder cancer for persons exposed to chlorinated drinking-water for 1-10, 11-20 or 21-30 years compared with those with no exposure to chlorinated drinking-water (Table 26C). The increased relative risk (OR = 1.8; 95% CI = 1.1-2.9) of bladder cancer for persons exposed to chlorinated drinking-water for more than 30 years was statistically significant. A trend of increasing risk with increasing duration of exposure was reported (trend P < 0.01), but no association was found between THM levels and bladder cancer (also see section 5.3.1.2). The authors reported that these results support the hypothesis that prolonged (i.e., more than 30 years) exposure to chlorinated drinking-water from surface water sources is associated with an increased risk of bladder cancer. However, incomplete characterization of the exposure profiles for water sources and disinfection, potential bias from non-participation and potential confounding bias from undetermined confounders tend to limit this interpretation. An increased risk of bladder cancer was seen among both smokers and non-smokers with more than 34 years of exposure to chlorinated water (Table 27B). Freedman et al. (1997) conducted a population-based case-control study in Washington County, Maryland (USA) to evaluate the association between the incidence of bladder cancer and use of chlorinated surface water (Table 26E). The study included 294 bladder cancer cases in white residents enumerated in a 1975 county census and reported to the county cancer registry between 1975 and 1992; 2326 white controls, frequency matched by age and gender, were randomly selected from the census. Duration of exposure to chlorinated surface water was based on length of residence in the census household before 1975, and relative risks were calculated using logistic regression methods adjusting for age, gender, tobacco use and urbanicity. Nearly all municipal sources in 1975 were supplied by surface waters that had been chlorinated for more than 30 years. Bladder cancer risk was found to be weakly associated with municipal water and duration of exposure to municipal water (for exposure to municipal water for more than 40 years, OR = 1.4; 95% CI = 0.7-2.9). The association was limited, however, to those who smoked cigarettes (Table 27C), primarily male smokers. Male smokers who had resided more than 40 years in an area with municipal water supplies had a 3-fold greater risk than male smokers who had not resided in an area with municipal water supplies or had resided in such an area for 20 years or less (OR = 3.2; 95% CI = 1.1-8.6). No increased relative risk was found for female smokers; in fact, relative risks for exposure to municipal water for 1-30 years were less than those for no exposure to municipal water (OR ranged from 0.4 to 0.6; 95% CI ranged from 0.1 to 2.9). A limitation of this study is that duration of exposure was based solely on place of residence in 1975. Also, there was no information on prior or subsequent domiciles, but the authors felt that because the population was relatively stable, residential mobility was not a major concern. No water consumption data were available, nor were data on DBPs available. As previously reported, however, chloroform analyses in 1975 found an average chloroform concentration of 107 g/litre in chlorinated surface waters used by the cohort (Wilkins & Comstock, 1981). Cantor et al. (1998) and Hildesheim et al. (1998) conducted a population-based case-control study in Iowa (USA) in 1986-1989 to evaluate cancer risks that may be associated with chlorinated water; results have been reported for bladder, colon and rectal cancer. Information about residential history, drinking-water source, beverage intake and other factors was combined with historical data from water utilities and measured THM levels to create indices of past exposure to chlorinated by-products. The bladder cancer study was composed of 1123 incident cases who were residents of Iowa, aged 40-85 years and diagnosed with histologically confirmed bladder cancer in 1986-1989 and 1983 randomly selected controls from driver's licence records and US Health Care Financing Administration listings for whom data relating to at least 70% of their lifetime drinking-water source were available. Of 1716 eligible bladder cancer cases, 85% participated. Adjusted odds ratios were determined using unconditional logistic regression analysis. Where appropriate, risks were adjusted for gender, age, cigarette smoking, years of education and employment in an occupation with elevated bladder cancer risk, including in the analysis persons with at least 70% of lifetime years with available information on drinking-water source. No increased risk was observed for men and women combined (Table 26D). Risk increased among men, with duration of chlorinated surface water exposure, duration of chlorinated groundwater exposure and duration of exposure to any chlorinated water source. For women, risks did not increase, and a protective effect for duration of exposure to chlorinated surface water was suggested, with OR values of less than unity; OR values ranged from 0.7 to 0.9, and 95% CIs ranged from 0.2 to 2.4. The increased relative risks in men were restricted to current smokers (Table 27D) and past smokers. No increased risk was found for men who had never smoked. Among non-smoking men and women, regardless of their previous smoking habit, there was no association between duration of exposure to chlorinated water and bladder cancer risk. Little or no association of risk was found for either total daily tapwater intake or intake of all beverages for men or women (Table 28A). Table 28. Bladder cancer risks associated with daily tapwater consumption Tapwater Odds 95% CI Comments Reference consumption ratio (litres/day) A. Males and females, Iowa Beverages from Cantor et al. <1.58 1.0 tapwater; adjusted for (1998) 1.58-<2.13 1.2 0.9-1.5 gender, age, study 2.13-2.85 1.3 1.0-1.6 period, education, >2.85 1.2 0.9-1.5 high-risk occupation, smoking B. Males and females, 10 areas Adjusted for age, Cantor et al. of the USA gender, smoking, (1985, 1987, <0.80 1.0 high-risk occupation, 1990) 0.81-1.12 1.1 0.9-1.3 population size, 1.13-1.44 1.1 1.0-1.3 usual residence 1.45-1.95 1.3 1.1-1.5 >1.95 1.4 1.2-1.7 Chloraminated water studies. In Massachusetts (USA), a number of towns have used surface water disinfected only with chlorine or chloramine since 1938, providing an opportunity to compare cancer risks between these two disinfectants. In a previous study, Zierler et al. (1986) found bladder cancer mortality to be weakly associated with residence at death in Massachusetts communities using chlorine for water disinfection; however, because of the concern that residence at time of death is a poor measure of historical exposure to a water disinfectant and likely resulted in misclassification bias, an analytical study (Zierler et al., 1988, 1990) was also conducted. Eligible for the study were all persons who were over 44 years of age at death and who died during 1978-1984 from bladder cancer, lung cancer, lymphoma, cardiovascular disease, cerebrovascular disease or chronic obstructive pulmonary disease while residing in 43 selected communities. Included were 614 persons who died of primary bladder cancer and 1074 individuals who died of other causes. Possible confounding bias by age, gender, smoking, occupation and socioeconomic status was controlled by multiple logistic regression. Detailed information on each person's residential history and possible confounding characteristics was obtained from survivors and census records. Analyses included a person's usual exposure (at least 50% of their residence since 1938 was in a community where surface water was disinfected by only one of the two disinfectants, either chlorine or chloramine) or lifetime exposure to water disinfected with only one of the two disinfectants. An association was found between bladder cancer mortality and both usual and lifetime exposure to chlorinated drinking-water. The bladder cancer mortality risk was higher for lifetime exposure than for usual exposure; a 60% increased risk of bladder cancer mortality (MOR = 1.6; 95% CI = 1.2-2.1) was found among lifetime residents of communities where only chlorinated surface water was used. The association is statistically significant, and the estimate of risk is precise (i.e., the CI is small). Bias is not likely present, but the magnitude of the association is not large and may be subject to residual confounding by unknown, unmeasured characteristics. The magnitude of the association found by Zierler et al. (1988, 1990), however, may be even larger than a 60% increased risk. Using only lymphatic cancers as the comparison group, the risk of bladder cancer mortality among lifetime consumers of chlorinated water was 3 times the risk for consumers of chloraminated drinking-water (MOR = 2.7; 95% CI = 1.7-4.3). This suggests that one or more causes of death in the comparison group may also be associated with water chlorination. McGeehin et al. (1993) also found that the risk of bladder cancer decreased with increasing duration of exposure to chloraminated surface water (trend P < 0.01). Persons who consumed chloraminated water for 21-40 years had a decreased, but not statistically significant, bladder cancer risk (OR = 0.7; 95% CI = 0.1-1.1). Those who consumed chloraminated water for more than 40 years also had a slightly decreased risk of cancer (OR = 0.6; 95% CI = 0.4-1.0). However, neither of these risks was statistically significant. McGeehin et al. (1993) reported that their results do not imply that chloraminated water conveys a protective effect, because there is no plausible biological explanation for suggesting that choramination inhibits neoplastic transformation of the bladder epithelium. Zierler et al. (1988, 1990) found that lifetime users of chloraminated surface water had a lower bladder cancer mortality risk than lifetime users of chlorinated surface water, and the findings of McGeehin et al. (1993), if real and not due to bias, provide additional evidence to support the conclusion that bladder cancer risks may be lower in persons using chloraminated drinking-water for long periods of time. Water and fluid consumption studies. Studies have reported both increased (Claude et al., 1986; Jensen et al., 1986) and decreased (Slattery et al., 1988) risk of bladder cancer associated with a high total fluid intake. Vena et al. (1993) conducted a case-control study that investigated the relationship between the incidence of bladder cancer and fluid intake and consumption of drinking-water. The study included 351 white males with histologically confirmed transitional cell carcinoma of the bladder and 855 white male controls with cancer of one of six other sites (oral cavity, oesophagus, stomach, colon, rectum and larynx). Study subjects were interviewed about diet and their total daily fluid intake of alcoholic beverages, bottled beverages, soda, milk, coffee, tea, all juices and glasses of tapwater. A dose-response relationship was found between daily intake of total liquids and risk of bladder cancer when a number of potential confounding factors were controlled (trend P < 0.001). For persons under the age of 65, up to a 6-fold increased risk of bladder cancer was found for those who drank more than 7 cups of fluids compared with those who drank 2-7 cups of total fluids daily (the lowest quartile studied). The OR ranged from 2.6 (95% CI = 1.2-5.7) for those who drank 8-10 cups of total fluids per day to 3.7 (95% CI = 1.7-8.2) for those who drank 11-13 cups and 6.3 (95% CI = 2.8-14.1) for those who drank 14-49 cups of total fluids daily. These ORs are adjusted for age, education, cigarette smoking, coffee, carotene and sodium by logistic regression. Statistically significant but slightly smaller risks were observed for those 65 and older. Total fluid consumption was divided into tapwater and non-tapwater consumption. Tapwater beverages included coffee, tea (hot and iced), 75% reconstituted orange juice, all other juices and glasses of water taken directly from the tap. A dose-response relationship was observed between increased tapwater consumption and bladder cancer (trend P < 0.001), but no statistically significant increased risk of bladder cancer, regardless of age, was found for persons consuming either 6-7 or 8-9 cups of tapwater per day compared with persons consuming 0-5 cups. Only in the highest group (10-39 cups of tapwater per day) was the association between tapwater consumption and bladder cancer statistically significant (OR = 2.6, 95% CI = 1.5-4.5, for those under the age of 65; OR = 3.0, 95% CI = 1.8-5.0, for those 65 and older). Increased bladder cancer risk was also found among persons under the age of 65 with the highest quartile of non-tapwater intake. The results provide additional hypotheses for further study, but limitations of this study preclude definitive conclusions regarding a potential water chlorination-cancer association. Over a third of potential cases were excluded for reasons including death (3%), refusal to be interviewed (24%) and extreme illness (11%). Selection bias would be present if the excluded cases were different from the participating cases in ways that are related to the exposure and outcome. Recall bias may also have been present, as bladder cancer cases were asked to recall usual dietary habits for the year before onset of cancer symptoms, whereas the controls, other cancer patients, were asked to recall dietary habits for the year before their interview for the study. Assessing fluid consumption in the year before onset of symptoms may not reflect historical or lifetime patterns, and increased consumption may occur during the early stages of bladder cancer. Since increased risks were seen primarily in total fluid consumption, it is also possible that total fluid consumption may be relevant in the pathogenesis of bladder cancer. Total daily fluid intake may be a marker for some unmeasured risk factor, and it is important to determine the biological relevance of increased fluid intake, independent of chlorinated tapwater, in the pathogenesis of bladder cancer (US EPA, 1994a). Since over 70% of the study population spent more than 90% of their lives using chlorinated surface water from public water supplies in western New York State, it is possible that the observed increased risks for bladder cancer may be associated with a high consumption of chlorinated tapwater. However, limited information was available with which to determine a study participant's duration of exposure to specific municipal water systems, and risks were not compared for populations using chlorinated surface water supplies. Cantor et al. (1987, 1990) conducted a further analysis of the national bladder cancer study to include beverage consumption information that was available for 5793 men and 1983 women. After correcting for age, smoking and other potential confounding characteristics, it was found that people who drank the most chlorinated tapwater had a bladder cancer relative risk about 40% higher than people who drank the least (Table 28B). When tapwater consumption was analysed separately for men and women, however, the association between water consumption and bladder cancer risk was statistically significant only among males. Bladder cancer risk was also evaluated for the combined effects of tapwater consumption and duration of chlorinated surface water use. No increased risk was associated with a high consumption of chlorinated drinking-water from surface water sources for less than 40 years. Increased bladder cancer risk (OR = 3.2; 95% CI = 1.2-8.7) was seen primarily in populations who had resided 60 or more years in areas served by chlorinated surface water and whose tapwater consumption was above the median of greater than 1.4 litres per day. Among non-smoking men, a risk gradient was apparent for those who consumed more than the population median of tapwater (Table 29), but a higher risk was also seen for non-smoking women who consumed less than the median amount. The increasingly smaller numbers of participants available for these subgroup analyses generally lead to statistically unstable estimates, making it difficult to evaluate trends in these data. In Colorado (USA) (McGeehin et al., 1993), the risk of bladder cancer (OR = 2.0; 95% CI = 1.1-2.8) was found to be elevated among persons who consumed more than five glasses of tapwater per day, and a dose-response trend was found ( P < 0.01). However, tapwater consumption appeared to be an independent risk factor for bladder cancer. There was no evidence of an increased relative risk of bladder cancer when both the volume of water consumed and years of exposure to chlorinated water were considered in the analysis. For example, similar estimates of risk were found for those who consumed more than five glasses of chlorinated water for fewer than 12 years (OR = 2.0; 95% CI = 0.8-4.7) and those who consumed more than five glasses of Table 29. Bladder cancer risks among non-smokers according to daily tapwater consumption and exposure to chlorinated surface water in 10 areas of the USAa Years at Odds ratio (95% CI)b residence with chlorinated water Tapwater consumption below 1.4 litres Tapwater consumption above 1.4 litres Males Females Males Females 0 1.0 1.0 1.0 1.0 1-19 1.6 (0.7-4.0) 1.2 (0.4-3.7) 0.8 (0.3-2.5) 1.7 (0.5-5.4) 20-39 1.0 (0.4-2.6) 1.5 (0.5-4.4) 2.1 (0.9-5.2) 1.8 (0.6-5.4) 40-59 0.7 (0.3-1.9) 2.1 (0.8-5.9) 2.5 (0.9-6.6) 1.8 (0.6-5.9) >60 1.3 (0.4-4.4) 4.3 (1.3-14.5) 3.7 (1.1-12.0) 3.6 (0.8-15.1) a From Cantor et al. (1987, 1990). b Adjusted for age, gender, smoking, high-risk occupation, population size, usual residence. chlorinated water for 12 years or more (OR = 2.4; 95% CI = 1.0-5.9). Similar to the study in western New York (Vena et al., 1993), tapwater consumption in Colorado was assessed for the year prior to diagnosis for cancer and controls and thus may not reflect historical patterns. King & Marrett (1996) also evaluated the combined effects of tapwater consumption and duration of exposure to chlorinated water (see also section 5.3.1.2). Overall, the pattern of risk estimates did not provide evidence for the interdependence of water consumption and years of exposure to THMs at levels above 49 g/litre to increase bladder cancer risks ( P for interaction = 0.775). Similar low, but not statistically significant, estimated relative risks are observed for those with less than 19 years' exposure to high THM levels, regardless of the volume of water they consumed. For persons with 20-34 years of exposure to THMs at levels above 49 g/litre, the estimated relative risk of bladder cancer was also similar for those who consumed less than 1.54 litres per day (OR = 1.7; 95% CI = 1.1-2.7) and those who consumed more than 2.08 litres per day (OR = 1.7; 95% CI = 1.1-2.7). For persons with 35 or more years of exposure to high THM levels, statistically significant risk estimates representing more than a doubling of estimated relative risk are observed for those who consumed between 1.54 and 2.08 litres of tapwater per day (OR = 2.6; 95% CI = 1.3-5.2) or more than 2.08 litres per day (OR = 2.3; 95% CI = 1.1-4.7). Cantor et al. (1998) studied tapwater consumption in Iowa (USA) and found little or no association of risk for either total daily tapwater intake or intake of all beverages for men or women (Table 28A). Colon cancer risk Cragle et al. (1985) investigated the relationship between water chlorination and colon cancer using 200 incident cases from seven hospitals and 407 hospital-based comparison subjects without evidence of cancer who had been North Carolina (USA) residents for at least 10 years. Comparison subjects were matched on age, race, gender, vital status and hospital to prevent potential confounding by these characteristics. Additional information on potential confounders, including alcohol consumption, genetic risk (number of first-degree relatives with cancer), diet, geographic region, urbanicity, education and number of pregnancies, was obtained by either mailed questionnaire or telephone interview. Water exposures were verified for each address and categorized as chlorinated or unchlorinated. Logistic regression analysis showed genetic risk, a combination of alcohol consumption and high-fat diet, and an interaction between age and chlorination to be positively associated with colo-rectal cancer. Risks for people who drank chlorinated water at their residences for 16 or more years were consistently higher than risks for those exposed to chlorinated water for less than 16 years, but a statistically significant association between water chlorination and colo-rectal cancer, controlling for possible confounding bias, was found only for those above age 60 (Table 30). For example, 70- to 79-year-old participants who drank chlorinated water for more than 15 years had twice the relative risk of colo-rectal cancer, but for 70- to 79-year old participants who drank chlorinated water for less than 16 years, the risk of colon cancer was only about 50% higher. Confusing the interpretation of the results is an apparent protective effect of chlorinated water for colo-rectal cancer found in age groups under age 50 (Table 30). For example, 40- to 49-year-old persons who drank chlorinated water for less than 16 years had about half the relative risk, and 20- to 29-year-old persons who drank chlorinated water for less than 16 years had about one-quarter the risk. These apparent protective effects may indicate lack of control for an important confounding characteristic. Table 30. Colon cancer risks associated with exposure to chlorinated water supplies in North Carolinaa Age Odds ratio (95% CI)b 1-15 years exposure >15 years exposure 20-29 0.2 (0.1-0.5) 0.5 (0.2-1.0) 30-39 0.4 (0.2-0.7) 0.6 (0.3-1.1) 40-49 0.6 (0.4-0.9) 0.8 (0.5-1.2) 50-59 0.9 (0.8-1.1) 0.9 (0.7-1.3) 60-69 1.2 (0.9-1.5) 1.4 (1.1-1.7) 70-79 1.5 (1.2-1.8) 2.2 (1.7-2.7) 80-89 1.8 (1.3-2.5) 3.4 (2.4-4.6) a From Cragle et al. (1985). b Adjusted for various confounders, including gender, race, diet, alcohol consumption, education, region and medical history of intestinal disorder. The colon cancer association in Wisconsin (USA) was further pursued in an interview-based study (Young et al., 1987, 1990) of incident cases of colon cancer and population-based comparison subjects where historical exposures to THMs were estimated. When water disinfection within only the most recent 10-year exposure period was considered, colon cancer cases were more likely supplied with chlorinated rather than with unchlorinated water (OR = 1.6; 95% CI = 1.0-2.4) and used municipal groundwater rather than private groundwater (OR = 1.7; 95% CI = 1.1-2.4). Increased risks were not found for use of chlorinated water or municipal groundwater for 20 or 30 years prior to diagnosis of cancer. THMs are not usually found in chlorinated municipal groundwaters in Wisconsin, but contaminants such as tetrachloroethylene, trichloroethylene and 1,1,1-trichloroethane have been found. Lawrence et al. (1984) studied the relationship of THMs and colo-rectal cancer in New York (USA) where THM exposure was higher than in Wisconsin (see section 5.3.1.2 for details about these studies). Hildesheim et al. (1998) conducted a population-based case-control study in Iowa (USA) in 1986-1989 to evaluate cancer risks that may be associated with chlorinated water; results have been reported for colon and rectal cancer. Information about residential history, drinking-water source, beverage intake and other factors was combined with historical data from water utilities and measured THM levels to create indices of past exposure to chlorinated by-products. The colon cancer study was composed of 560 incident cases who were residents of Iowa during March-December 1987, aged 40-85 years and with histological confirmation and 2434 age and gender frequency matched controls for whom water exposure information was available for at least 70% of their lifetime. Of the 801 eligible colon cancer cases, 685 (86%) participated; of these cases, 560 (82%) had sufficient information about water exposures. Unconditional multiple logistic regression analysis was used to estimate odds ratios for risks associated with chlorinated surface water and groundwater, THM exposure and tapwater consumption, while adjusting for potentially confounding factors. For colon cancer and subsites, no increase in risk was associated with duration of exposure to chlorinated surface water or chlorinated groundwater (Table 31A). A slight decrease in colon cancer risk was found with increased tapwater consumption. Those who drank 2.9 litres or more of tapwater daily had a 25% reduced risk of colon cancer compared with those who drank less than 1.5 litres per day. Rectal cancer risk Hildesheim et al. (1998) conducted a population-based case-control study in Iowa (USA) in 1986-1989 to evaluate cancer risks that may be associated with chlorinated water; results have been reported for colon and rectal cancer. Information about residential history, drinking-water source, beverage intake and other factors was combined with historical data from water utilities and measured THM levels to create indices of past exposure to chlorinated by-products. The study was composed of 537 incident rectal cancer cases aged 40-85 years who were residents of Iowa from January 1986 to December 1988 and whose cancer was confirmed histologically and 2434 age and gender frequency matched controls for whom water exposure information was available for at least 70% of their lifetime. Of the 761 eligible rectal cancer cases, 655 (86%) participated; sufficient information about water exposures was available for 537 (82%) of these 655 cases. Unconditional multiple logistic regression analysis was used to estimate odds ratios for risks associated with chlorinated surface water, chlorinated groundwater, THM exposure and tapwater consumption, while adjusting for potentially confounding factors. An increasing rectal cancer risk was associated with both increasing cumulative THM Table 31. Colo-rectal cancer risks associated with exposures to chlorinated water supplies in Iowa (USA)a Years of exposure Odds 95% CI Comments to chlorinated ratio surface water A. Colon cancer risks Adjusted for age, 0 1.0 gender 1-19 1.0 0.8-1.3 20-39 1.0 0.7-1.5 40-59 1.2 0.8-1.8 >60 0.8 0.4-1.7 B. Rectal cancer risks Adjusted for age, 0 1.0 gender 1-19 1.1 0.8-1.4 20-39 1.6 1.1-2.2 40-59 1.6 1.0-2.6 >60 2.6 1.4-5.0 a From Hildesheim et al. (1998). exposure and duration of exposure to chlorinated surface water (Table 31B). However, the amount of tapwater consumed did not confound the risk, as the authors reported that little association (no data provided) was found between water consumption and rectal cancer after adjustment for age and gender. Larger relative risks for rectal cancer were found among persons with low dietary fibre intake and longer-duration exposure to chlorinated surface water source compared with persons with high-fibre diets and no exposure to chlorinated surface water. Pancreatic cancer risk A population-based case-control study (Ijsselmuiden et al., 1992) was conducted in Washington County, Maryland (USA), the same area in which an earlier cohort study had been conducted by Wilkins & Comstock (1981). Included in the study were 101 residents who were identified by the county cancer registry with a diagnosis of pancreatic cancer from July 1975 to December 1989 and 206 controls randomly chosen from the county population defined by a specially conducted census in 1975. Drinking-water source obtained from the 1975 census was used to assess exposure. Multivariate analysis found an increased risk of pancreatic cancer (RR = 2.2; 95% CI = 1.2-4.1) associated with chlorinated municipal surface water after adjusting for cigarette smoking; however, the authors recommended caution in the interpretation of this finding because of limitations in the assessment of exposure for study participants. Exposure was assessed at only one point in time, 1975, and this may not accurately reflect long-term exposure to chlorinated water. No increased risk of pancreatic cancer (RR = 0.8; 95% CI = 0.4-1.5) was found by Wilkins & Comstock (1981) in an earlier cohort study in the same county. Ijsselmuiden et al. (1992) did not report the reason for studying pancreatic cancer rather than evaluating further the previous findings of Wilkins & Comstock (1981) of increased, but not statistically significant, risks for bladder, liver or kidney cancer. Brain cancer risk In an abstract, Cantor et al. (1996) described the results of a population-based case-control study of 375 incident brain cancer patients, diagnosed in 1984-1987, and 2434 controls in Iowa. After controlling for age, farm occupation and other potential confounding characteristics, brain cancer risk among men, but not women, was associated with increased duration of exposure to chlorinated surface water. The risk was greatest for over 40 years of exposure (OR = 2.4; no CI reported) compared with 20-39 years of exposure (OR = 1.8) and 1-19 years of exposure (OR = 1.3). Historical exposures to chlorination by-products in drinking-water were also estimated from recent measures of THMs, other water quality data and information from study participants, but results of these analyses were not presented. 5.2.1.3 Meta-analysis of cancer studies Meta-analysis is the application of quantitative techniques to literature reviewing. Two complementary approaches may be taken (US EPA, 1994a). One is an aggregative approach to summarize the compiled research on a given topic. This summary typically provides a measure of overall statistical significance or a consolidated estimate of effect, such as a relative risk. In practice, "when the world's literature on a topic is declared to be statistically significant," results of this type of meta-analysis are usually interpreted as an indication that research should cease and action should begin (US EPA, 1994a). The other approach to meta-analysis is an analytical or explanatory approach (Greenland, 1994), in which the goal is to see whether differences among the studies can explain differences among their results. A formal analysis of the explanatory type might take the form of a meta-regression (Greenland, 1987), in which the dependent variable is the measure of effect, as estimated by each study, and the independent variables are the potentially explanatory factors that might yield higher or lower estimates of effect. The strength of explanatory meta-analysis is that it produces an enhanced understanding and appreciation of the strengths and weaknesses of the studies that have been conducted on the selected topic (US EPA, 1994a). In the case of chlorinated drinking-water and cancer, a meta-analysis that used the aggregative approach was published by Morris et al. (1992). The authors conducted significance tests for several cancers and reported two, bladder cancer and rectal cancer, as statistically significant. Emphasized were the summary estimates of effect: pooled RRs of 1.2 (95% CI = 1.1-1.2) for bladder cancer and 1.4 (95% CI = 1.1-1.9) for rectal cancer. In conjunction with crude data on population exposure to chlorinated surface water, an attributable risk was estimated by the authors, suggesting that about 4200 cases (9%) of bladder cancer per year and 6500 cases (18%) of rectal cancer per year may be associated with consumption of chlorinated surface water in the USA. The meta-analysis and its quantitative estimate of risk have caused considerable controversy because there are significant differences in the design of the 10 individual epidemiological studies included (IARC, 1991; Risch et al., 1992; Craun et al., 1993; Murphy, 1993; Bailar, 1995). Reported quality scores for the individual studies were low (43-78 of a possible 100), and their study populations, research methods and results are not homogeneous. Several of the included studies were well designed and attempted to assess historical exposures to chlorinated water and possible confounding characteristics, but most did not adequately assess historical exposures and confounding bias. Only three studies adjusted or matched for smoking (one bladder cancer study, one colon cancer study and a cohort study of bladder, colon and rectal cancer), and only one study considered diet as a potential confounder. In four studies, a single reported address on a death certificate was used to assess the study participants' exposure to chlorinated water. It is important to ask whether the meta-analysis should have restricted the analysis to those studies that were more homogeneous or those with adequate exposure assessments and control of confounding bias. Bailar (1995) questioned the summary findings for bladder cancer based on his assessment of the weaknesses of the individual studies and concluded that "bias could well explain the whole of the apparently positive findings." A recently completed formal evaluation of the meta-analysis by Poole (1997) concluded that studies should not have been combined into aggregate estimates of relative risk or used as the basis for national attributable risk estimates. 5.2.1.4 Summary of results of cancer studies Various types of epidemiological studies, primarily in the USA, have attempted to assess the cancer risks associated with chlorinated water systems. Water disinfected with ozone and chlorine dioxide has not been studied for cancer associations, but chloraminated drinking-water was considered in two studies. Many, but not all, ecological studies reported associations between chlorinated water and cancer incidence or mortality and helped develop hypotheses for further study. Analytical studies reported small relative risks for colon and bladder cancer incidence for populations consuming chlorinated drinking-water for long periods of time. Because of probable bias, interpretation of observed associations is severely limited in case-control studies where information was not obtained from interviews and residence histories. Interview-based case-control epidemiological studies provide a basis for evaluating the potential cancer risk that may be associated with chlorinated drinking-water. Based on Monson's (1980) guide to interpreting the strength of an association, a weak to moderate epidemiological association was found between water chlorination and colon cancer incidence among an elderly population in North Carolina (USA). However, a moderate to strong protective effect was also found among persons 20-49 years of age, confusing the interpretation of these results. The higher risks for those above age 70 in North Carolina suggest that the association may be evident only after a very long duration of exposure to chlorinated surface water. Colon cancer incidence was weakly associated with the use of chlorinated drinking-water for the most recent 10-year period in Wisconsin (USA), but no association was found when 20 or 30 years of exposure were considered. In Iowa (USA), no association was found between colon cancer or any subsites and duration of exposure to chlorinated surface water. In a national study of 8764 persons in the USA, no overall association was found between chlorinated drinking-water and bladder cancer risk. A moderate association between chlorinated surface water and bladder cancer incidence was observed among an otherwise low-risk population of non-smokers that had received chlorinated surface water for 60 or more years, but this analysis included only 123 persons. In the Iowa portion of this study, moderate to strong associations were found for smokers and non-smokers with at least 30 years of exposure to chlorinated water. Interview case-control studies found a moderate risk of bladder cancer incidence associated with more than 30 years of exposure to chlorinated surface water in Colorado (USA), a weak to moderate risk of bladder cancer incidence associated with more than 35 years of exposure to chlorinated surface water in Ontario (Canada), a weak to moderate risk associated with more than 59 years of exposure to chlorinated surface water in Iowa (USA), and a weak association with more than 40 years of exposure to chlorinated surface water in Washington County, Maryland (USA). Inconsistencies, however, were observed in risks for non-smokers and smokers and in risks for women and men. In New York State (USA), a dose-response relationship was observed between daily intake of total liquids and risk of bladder cancer and between increased tapwater consumption and risk of bladder cancer. Since increased risks were seen primarily in total fluid consumption, it is possible that total fluid consumption may be relevant in the pathogenesis of bladder cancer. Total daily fluid intake may be a marker for some unmeasured risk factor, and it is important to determine the biological relevance of increased fluid intake, independent of chlorinated tapwater, in the pathogenesis of bladder cancer. Several studies found the risk of bladder cancer to be elevated among persons who consumed more tapwater per day, but increased tapwater consumption appears to be an independent risk factor for bladder cancer. There was no evidence of an increased relative risk of bladder cancer when both the volume of water consumed and duration of exposure to chlorinated water or THMs were considered in the analysis. In Massachusetts (USA), an increased risk of bladder cancer mortality was observed in a population receiving chlorinated surface water compared with a population receiving chloraminated surface water. These associations are weak to moderate in strength, depending upon the diseases used for comparison, and also considered long exposures, over 40 years' duration. A decreased risk of bladder cancer incidence was also associated with a similar duration of exposure to chloraminated surface water in Colorado (USA), but investigators felt that the results do not imply that chloraminated water conveys a protective effect because there is no plausible biological explanation for suggesting that choramination inhibits neoplastic transformation of the bladder epithelium. A case-control study reported a large increased risk of rectal cancer among those with long duration of exposure to chlorinated water, but two cohort studies did not find an increased relative risk. A single case-control study reported a moderate to strong risk of pancreatic cancer associated with chlorinated surface water in Washington County, Maryland (USA), but the interpretation of this study is hampered because exposure was assessed at only one point in time, and this may not accurately reflect long-term exposure to chlorinated water. Preliminary results from a study in Iowa (USA) found that a moderate risk of brain cancer among men, but not women, was associated with increased duration of exposure to chlorinated surface water, especially for over 40 years of exposure. Additional details of this study are required to allow these conclusions to be evaluated. A weak increased risk of lung cancer incidence was also seen for users of chlorinated surface water in Iowa. A controversial meta-analysis study, which statistically combined the results of 10 previously published epidemiological studies, reported a small pooled increased relative risk for bladder and rectal cancer but not for colon cancer. This meta-analysis, however, included a number of low-quality studies with likely bias. Current evidence from epidemiological studies is insufficient to allow a causal relationship between the use of chlorinated drinking-water and the incidence of bladder cancer to be established. Several studies reported weak to moderate associations of long-duration exposure to chlorinated water and bladder cancer, but risks have differed between smokers and non-smokers in several studies. Inconsistent risks have also been seen when gender and water consumption were considered. For colon cancer, the epidemiological data appear to be equivocal and inconclusive. For rectal cancer, insufficient data are available with which to evaluate the moderate associations observed in one study. Similarly, single studies of reported associations for pancreatic, lung, brain and breast provide insufficient data. 5.2.2 Epidemiological studies of cardiovascular disease and disinfected drinking-water Several epidemiological studies have evaluated possible risks of cardiovascular disease associated with the chlorination of drinking-water. The cohort study (Wilkins & Comstock, 1981) of 31 000 residents of Washington County, Maryland (USA) found a slightly increased, but not statistically significant, risk (RR = 1.1; 95% CI = 1.0-1.3) of death due to arteriosclerotic heart disease in residents exposed to chlorinated drinking-water from a surface water and springs (average chloroform level was reported as 107 g/litre) compared with residents of towns where unchlorinated well-water was used. A standardized mortality study (Zierler et al., 1986) found little difference in the patterns of 35 539 and 166 433 deaths (age of death was at least 45 years) due to cerebrovascular and cardiovascular disease, respectively, during 1969-1983 in 43 Massachusetts (USA) communities with water supplies disinfected with either chlorine or chloramine as compared with all deaths reported in Massachusetts due to these causes. The mortality rate in these selected communities with chlorinated drinking-water was slightly higher than expected for cerebrovascular disease (SMR = 108; 95% CI = 106-109) and cardiovascular disease (SMR = 104; 95% CI = 104-105). The mortality rate in selected communities with chloraminated drinking-water was slightly lower than expected for cerebrovascular disease (SMR = 86; 95% CI = 85-88) and about the same as expected for cardiovascular disease (SMR = 101; 95% CI = 100-101). As noted in section 5.2.1.2, the serious potential for exposure and disease misclassification bias in the Massachusetts study limits the interpretation of these findings. A cross-sectional study (Zeighami et al., 1990a,b) of 1520 adult residents, aged 40-70 years, in 46 Wisconsin (USA) communities was conducted to determine whether the hardness or chlorination of drinking-water affects serum lipids. The water for the communities contained total hardness of less than or equal to either 80 or 200 mg of calcium carbonate per litre; 858 participants (59% female) resided in 24 communities that provided chlorinated water, and 662 (55% female) resided in 22 communities that did not disinfect water. An age-gender-stratified sampling technique was used to choose a single participant from each eligible household, and a questionnaire was administered to obtain data on occupation, health history, medications, dietary history, water use, water supply and other basic demographic information. Among women who resided in communities with chlorinated, hard drinking-water, mean serum cholesterol levels were found to be higher (249.6 mg/dl, standard error [SE] = 6.4) than for women who resided in unchlorinated, hard-water communities (235.3 mg/dl, SE = 6.4). Among women who resided in communities with chlorinated, soft drinking-water, mean serum cholesterol levels were also found to be higher (248.0 mg/dl, SE = 6.2) than in women residing in unchlorinated, soft-water communities (239.7 mg/dl, SE = 6.3). Mean serum cholesterol levels were also higher for men in chlorinated communities, but the differences in mean cholesterol levels between chlorinated and unchlorinated communities were smaller and not statistically significant. The age-specific and overall risks of elevated serum cholesterol levels (>270 mg/dl) in communities with chlorinated drinking-water were also evaluated for men and women. Statistically significant increased risks of elevated serum cholesterol levels were found among women but not men and primarily in women aged 50-59 (OR = 3.1; 95% CI = 1.5-6.7). Mean levels of LDL cholesterol in men and women had a similar pattern to total cholesterol. However, mean levels of HDL cholesterol were nearly identical in the chlorinated and unchlorinated communities for each gender, and the implication for increased cardiovascular disease risk in communities with chlorinated water remains unclear. Caution is urged in the interpretation of the results of this study. It is possible that an undetermined confounding characteristic due to lifestyle differences may be responsible for the observed association in the chlorinated communities. The relationship between consumption of chlorinated drinking-water in the home and serum lipids was also evaluated in a cohort of 2070 white women, aged 65-93, participating in a study of osteoporotic fractures in western Pennsylvania (USA) (Riley et al., 1995). Mean serum cholesterol levels (247 mg/dl, SD = 41.3) in 1869 women using chlorinated water sources were similar to levels (246 mg/dl, SD = 41.9) in 201 women using unchlorinated water sources. Women with the largest cumulative exposure to chlorinated water had higher serum cholesterol levels (247 mg/dl, SD = 41.3) than women with unchlorinated water (241 mg/dl, SD = 0.7), but the difference was not statistically significant. There was no evidence that increasing the duration of exposure to chlorinated water influenced LDL or HDL cholesterol, triglycerides or apolipoproteins. In this cohort, women exposed to chlorinated water sources tended to smoke and drink more than women not exposed to chlorinated water sources, suggesting that the reported association in Wisconsin may be due primarily to inadequate control of lifestyle characteristics differentially distributed across chlorinated exposure groups. 5.2.2.1 Summary of results of cardiovascular studies A mortality study showed little difference between patterns of death due to cardiovascular and cerebrovascular disease in the general population of Massachusetts (USA) and among residents of communities using either chlorine or chloramine . A cohort study in Washington County, Maryland (USA) found a slightly increased, but not statistically significant, risk of death due to arteriosclerotic heart disease in residents exposed to chlorinated drinking-water. A cross-sectional epidemiological study in Wisconsin (USA) found higher serum cholesterol in those exposed to chlorinated drinking-water than in those not exposed, but the association was confined to women aged 50-59. Mean serum lipids and lipoproteins were found to be similar in elderly white women exposed to chlorinated and unchlorinated drinking-water in western Pennsylvania (USA). There is inadequate evidence from epidemiological studies that chlorinated or chloraminated drinking-water increases cardiovascular disease risks. 5.2.3 Epidemiological studies of adverse reproductive/developmental outcomes and disinfected drinking-water After adjustment for potential confounders, Aschengrau et al. (1989) found a statistically significant increase in the frequency of spontaneous abortion in Massachusetts (USA) communities that used surface water sources compared with those that used groundwater sources. No measures of DBPs were available for the communities. When surface water sources are chlorinated, DBPs are typically higher than when groundwater sources are chlorinated, but levels of other water parameters also differ between surface water and groundwater sources. Aschengrau et al. (1993) also conducted a case-control study of late adverse pregnancy outcomes and water quality in Massachusetts community water systems. Among women who delivered infants during August 1977 and March 1980 at Brigham and Women's Hospital, various water quality indices were compared for 1039 congenital anomaly cases, 77 stillbirth cases, 55 neonatal deaths and 1177 controls. Risk of neonatal death or all congenital anomalies was not found to be increased in women exposed to chlorinated surface water supplies compared with chloraminated water supplies. Stillbirth risk was associated with chlorinated water supplies; however, after adjustment for appropriate confounding characteristics, the risk was not statistically significant (OR = 2.6; 95% CI = 0.9-7.5). Water quality measures were available for trace metals, but no water quality measures were available with which to assess risks associated with THMs or other DBPs. A population-based case-control study of miscarriage, preterm delivery and low birth weight was conducted in three counties in central North Carolina (USA) (Savitz et al., 1995). Preterm deliveries (<37 weeks completed gestation) and low birth weight infants (<2500 g) were identified at hospitals with virtually all births to residents of Orange and Durham counties from September 1988 to August 1989 and Alamance County from September 1988 to April 1991. About 50% of eligible live births were both preterm and low birth weight. All medically treated miscarriages among women in Alamance County from September 1988 to August 1991 were also identified for study. Full-term, normal-weight births immediately following a preterm or low birth weight delivery were selected as controls. Race and hospital were controlled for in the analysis. Telephone interviews were used to obtain information on a number of potential risk factors, including age, race, education, marital status, income, pregnancy history, tobacco and alcohol use, prenatal care, employment and psychological stress. Questions about drinking-water sources at home, including bottled water, and amount of water consumed around the time of pregnancy were also asked. Only 62-71% participated; the lowest response rate was for miscarriage cases. The analysis, which considered water consumption and THMs, was further restricted to women served by public water sources who reported drinking one or more glasses of water per day, and this limited the analysis to 70% of those participating. Selection bias is of concern in these studies, as is recall bias, where cases are more likely to more accurately recall previous exposures. It was found that water source was not related to miscarriage, preterm delivery or low birth weight. Risk of miscarriage was slightly increased among women who used bottled water compared with those who used private wells, but the risk was not statistically significant (OR = 1.6; 95% CI = 0.6-4.3). A cross-sectional study (Kanitz et al., 1996) was conducted of 548 births at Galliera Hospital in Genoa and 128 births at Chiavari Hospital in Chiavari (Italy) during 1988-1989 to mothers residing in each city. Women in Genoa were exposed to filtered water disinfected with chlorine dioxide (Brugneto River wells, reservoir and surface water) and/or chlorine (Val Noci reservoir). Women residing in Chiavari used untreated well-water. Assignment to a water source and type of disinfectant was based on the mother's address (undisinfected well-water, chlorine, chlorine dioxide or both). Municipal records were used to determine family income, and hospital records were used to obtain information about mother's age, smoking, alcohol consumption and education level and birth outcomes -- low birth weight (<2500 g), preterm delivery (<37 weeks), body length (<49.5 cm), cranial circumference (<35 cm) and neonatal jaundice. Neonatal jaundice was almost twice as likely (OR = 1.7; 95% CL = 1.1-3.1) in infants whose mothers resided in the area where drinking-water from surface water sources was disinfected with chlorine dioxide as in infants whose mothers used undisinfected well-water. No increased risk for neonatal jaundice was found for infants whose mothers lived in an area using chlorinated surface water. Large increased risks of smaller cranial circumference and body length were associated with drinking-water from surface water sources disinfected with chlorine or chlorine dioxide. Infants born to mothers residing in areas where surface water was disinfected with chlorine or chlorine dioxide had smaller cranial circumference (OR = 3.5; 95% CI = 2.1-8.5 for chlorine vs. untreated well-water; OR = 2.2; 95% CI = 1.4-3.9 for chlorine dioxide vs. untreated well water) and smaller body length (OR = 2.0; 95% CI = 1.2-3.3 for chlorine dioxide vs. untreated well-water; OR = 2.3; 95% CI = 1.3-4.2 for chlorine vs. untreated well-water). Risks of low birth weight infants were also increased for mothers residing in areas using water disinfected with chlorine and chlorine dioxide, but these associations were not statistically significant. For preterm delivery, small and not statistically significant increased risks were found among mothers residing in the area using chlorine dioxide. This study suggests possible risks associated with surface water disinfected with chlorine or chlorine dioxide, but the results should be interpreted very cautiously (US EPA, 1997). THM levels were low in both chlorinated water (8-16 g/litre) and chlorine dioxide-disinfected water (1-3 g/litre). No information was collected to assess the mothers' water consumption or nutritional habits, and the age distribution of the mothers was not considered. It is important to determine whether municipal or bottled water was consumed by the mothers and how much water was consumed. If mothers routinely consumed untreated municipal well-water but did not consume disinfected municipal surface water, drinking bottled water instead, the observed relative risks would not be associated with the disinfection of surface water. In addition, there are concerns about incomplete ascertainment of births and whether the population may be different in respects other than the studied water system differences. On the other hand, if the observed associations with water source and disinfection are not spurious, a question is raised about what water contaminants may be responsible. Exposures to surface water and groundwater sources are compared in this study, and no information is presented about other possible water quality differences. Preliminary analyses were available for a cross-sectional study (Nuckols et al., 1995) of births to women in Northglenn, Colorado (USA) who were exposed to chlorinated water from Stanley Lake with high THM levels (32-72 g/litre) and to women in Westminster who were exposed to chloraminated water from Stanley Lake with low THM levels (<20 g/litre). Lower, but not statistically significant, relative risks for low infant birth weight and preterm delivery were found in the water district using chlorine. Further analysis in the chlorinated water system found increased, but not statistically significant, relative risks for low birth weight infants and preterm delivery in areas where the THMs were higher. These results must be interpreted with caution because very limited information was provided about the epidemiological methods. Instead, the article focused on water distribution system quality modelling and the use of geographic information systems. 5.2.3.1 Summary of results of reproductive/developmental studies Results of several exploratory epidemiological studies of adverse reproductive effects/developmental outcomes and chlorinated water should be cautiously interpreted because of limitations of study design and likely bias. 5.3 Epidemiological associations between disinfectant by-products and adverse health outcomes In this section, studies of specific DBPs are reviewed. In some studies, both water source and disinfectant type were considered in addition to the specific by-products; if this is the case, the details of the study design and any limitations are reported in section 5.2. 5.3.1 Epidemiological studies of cancer and disinfectant by-products 5.3.1.1 Cancer associations in ecological studies 1) Volatile by-products Cantor et al. (1978) studied the relationship between THM levels and age-standardized cancer mortality for 1968-1971 of white men and women in urban counties in the USA. THM concentrations in drinking-water were estimated at the county level from data obtained in two national surveys of water supplies conducted by the US EPA. The analysis took into account the median number of school years completed by county inhabitants over age 25, the foreign-born and native population of the county, the change in county population from 1950 to 1970, the percentage of each county that is considered urban, the percentage of the county work force engaged in all manufacturing industries, and the geographic region of the USA. Multivariate regression analysis found the variability among these counties for gender-specific and cancer-site-specific mortality rates, and the residual mortality rates were then correlated directly with estimated THM exposure for those 76 counties in which 50% or more of the population was served by the sampled water supplies. Among both men and women, a statistically significant positive correlation was found between non-chloroform THM levels and bladder cancer mortality. No association between THM levels and colon cancer mortality rates was observed after controlling for the ethnicity of the population. Hogan et al. (1979) studied county cancer mortality data for an earlier period and county chloroform levels estimated from the same EPA surveys. Multivariate regression analysis of the county cancer mortality rates included, as independent variables, estimated exposure to chloroform concentrations, 1960 county population, county population density, percentage of county that is urban, percentage of county population that is non-white, percentage of county population that is foreign-born, median number of school years completed by county residents over age 25, median family income of county and percentage of county work force engaged in manufacturing. The results suggested that county cancer mortality rates for the rectum, bladder and possibly large intestine increased with increased levels of chloroform in drinking-water supplies. McCabe (1975) found that age-adjusted total cancer mortality rates correlated positively with estimated chloroform concentrations in 80 US cities but included no attempt to control for potential confounding bias on a group or aggregate level. Carlo & Mettlin (1980) studied 4255 incident cases of oesophageal, stomach, colon, rectal, bladder and pancreatic cancers reported through the New York State Tumor Registry for Erie County, New York (USA) between 1973 and 1976. Age-adjusted incidence rates were calculated by census tract, and levels of THMs were estimated from a single water survey in July 1978. Statistically significant positive associations were found between consumption of chlorinated drinking-water from surface water sources and oesophageal and pancreatic cancer and between THM levels and pancreatic cancer in white males. The authors placed little credence on these findings, noting that the pancreatic cancer-THMs relationship was found only in one gender-race subgroup, the range of THM concentrations in the study was narrow (the largest variation was only 71 g/litre) and no data were available with which to estimate historical trends in THM levels. Tuthill & Moore (1980) studied cancer mortality rates from 1969 to 1976 in Massachusetts (USA) communities supplied by surface water. Chlorination exposure data were assessed by considering past chlorine dosage, recent THM levels and recent chlorine dosage. Stomach and rectal cancer mortality rates were associated with recent THM levels and recent chlorine dosage, but not with past chlorine dosage in the communities. However, when regression models included migration patterns and ethnicity, none of the associations was statistically significant. Wigle et al. (1986) studied selected contaminants in drinking-water and cancer risks in Canadian cities with populations of at least 10 000. Water quality data from three national surveys of urban drinking-water supplies, demographic data and age-standardized cancer mortality rates for 1973-1979 were analysed by multivariate regression techniques. No statistically significant associations were found between chlorine dosage and risk of death with any disease category. When chlorine dosage was replaced in the model by TOC, a statistically significant association was found between this variable and cancer of the large intestine among males but not females. There were no statistically significant associations when chlorine dosage was replaced by THM, chloroform or non-chloroform THM levels. A recent ecological analysis of cancer cases reported to the New Jersey (USA) cancer registry in 1979-1990 found no association between bladder and rectal cancer incidence and THM or BDCM levels in public drinking-water systems (Savin & Cohn, 1996). Exposure was based on address at the time of cancer diagnosis and the water quality of the public water system from monitoring conducted during the 1980s. These ecological studies used limited information about current THM levels in drinking-water to estimate exposures for a census tract, county or community, and these exposures may not represent long-term exposures or be relevant for the population whose cancer statistics were studied. Available demographic characteristics were included as group variables to assess or control for confounding bias, but these likely have limited value in controlling for such bias. Because of the ecological design of the study, these results cannot be easily interpreted. Even if the exposure information were accurately assessed, it cannot be determined whether the observed associations were a result of exposure to chloroform, THM levels or other DBPs or were confounded by characteristics that were not assessed (e.g., cigarette smoking) or incorrectly assessed by using available demographic data. Ecological studies where no associations were found suffer from similar limitations. 2) Mutagenicity Several ecological studies have considered the mutagenic activity of drinking-water, as determined by the Salmonella/microsome assay. These mutagenicity tests assess the non-volatile, acid/neutral fraction of chlorinated organic material in a concentrated water sample. A mutagen of particular concern is MX, a potent mutagen, as measured by strain S. typhimurium TA100, which may be responsible for up to 57% of the mutagenicity in chlorinated drinking-water (Meier et al., 1986, 1987b). High levels of mutagenic activity have been observed in Finnish chlorinated drinking-water, and Koivusalo et al. (1994a,b, 1995) investigated the relationship between mutagenic activity in drinking-water and gastrointestinal, urinary tract and other cancers in 56 Finnish municipalities. Included in the analysis were cases of bladder, kidney, stomach, colon, rectum, liver, pancreas and soft tissue cancer, leukaemia (acute, chronic myeloid and chronic lymphatic), Hodgkin's disease and non-Hodgkin's lymphoma obtained from the population-based Finnish Cancer Registry for two periods, 1966-1976 and 1977-1989. The Cancer Registry includes virtually all cancer cases in the country. On an ecological level, information was also obtained for several potential confounding characteristics -- social class, urban living, time period, migration and exposures from the chemical industry. Koivusalo et al. (1994b) discussed the methodology used to estimate the mutagenicity of drinking-water and assess past exposures for the epidemiological studies. Exposure was assessed at the ecological, not the individual, level and was based on the proportion of population served by municipal water and the estimated mutagenicity of the water. Previous drinking-water mutagenic activity was estimated for each community on the basis of an equation derived by Vartiainen et al. (1988), yielding an estimated drinking-water mutagenicity level in net revertants per litre for each community for each of two specific periods, 1955 and 1970. The equation used available historical information on raw water quality (TOC, ammonia, permanganate test for oxidizable organic matter, etc.), pre- and post-chlorine dosages, and water treatment practices obtained by a questionnaire sent to the municipalities. Vartiainen et al. (1988) and Koivusalo et al. (1994b) reported a high correlation between estimated and measured drinking-water mutagenicity after comparing the results of assays of drinking-water mutagenicity in 1985 and 1987 with estimated mutagenicity from their equation. Observed and expected numbers of cancer cases in each municipality were compared by sex, broad age group and time period, and cancer risk was adjusted for social class as a surrogate for lifestyle and smoking habits. For all 56 municipalities, no statistically significant increases in risks of cancer of the kidney, stomach, colon, rectum, liver, pancreas and soft tissue, leukaemia (acute, chronic myeloid and chronic lymphatic) and non-Hodgkin's lymphoma were found among those exposed to a typically mutagenic drinking-water (3000 net revertants per litre) when adjusted for age, sex, period, main cities and social class. The risks of bladder cancer (RR = 1.2; 95% CI = 1.1-1.3) and Hodgkin's disease (RR = 1.2; 95% CI = 1.0-1.4) were small but statistically significant. When this analysis was restricted to those 34 municipalities with mutagenic drinking-water, risks of non-Hodgkin's lymphoma, Hodgkin's disease and cancer of the liver, pancreas, kidney, stomach and bladder were all found to be slightly increased (RRs ranged from 1.1 to 1.3). As this is an ecological study, the small increased relative risks that were observed preclude definitive conclusions. 5.3.1.2 Cancer associations in analytical studies 1) Bladder cancer risk In Colorado (USA), where an increased risk of bladder cancer was found for persons exposed to chlorinated drinking-water for more than 30 years (OR = 1.8; 95% CI = 1.1-2.9), no association was found between THM levels and bladder cancer (McGeehin et al., 1993). The mean 1989 levels for THMs and residual chlorine for each water system were multiplied by the number of years each study participant was exposed to that system and summed to compute lifetime exposure indices for each water quality indicator. Higher risks were found when the cumulative exposure index for THMs was less than 200 or greater than 600 g/litre-years, but not when it was 201-600 g/litre-years. No statistically significant trends were found; risks did not increase with increased cumulative THM exposure. In logistic regression models that included water source and disinfection variables and controlled for years of exposure to chlorinated water, THMs assessed by this exposure index were not associated with bladder cancer. Interpretation of these results is limited, however, because only 61% of THM levels and 55% of residual chlorine levels could be determined for use in assessing exposures. In Ontario (Canada), individual information about water consumption and exposure to chlorinated water was supplemented by water treatment data (e.g., area served, water source and characteristics, and treatment practices for years of operation between 1950 and 1990) collected from a survey of historical treatment practices at water supplies in the study area (King & Marrett, 1996). For each treatment facility, water treatment information was obtained for an average day in August in 5-year intervals, and that observation was used to represent water characteristics for the years surrounding that date. The water supply for each participant was characterized by the estimated annual maximum THM levels for each water facility. Historical THM levels were estimated using a model developed to predict the THM level in treated water from characteristics of the treatment process. The model was developed from water quality measurements recorded by the Ontario Drinking Water Surveillance Program between 1988 and 1992 for 114 water treatment facilities. Predicted THM levels were compared with those observed in 1986-1987, and the correlation between observed and predicted values was 0.76. When observed and predicted THM levels were considered as either above or below 50 g/litre, the model predicted values with a sensitivity of 84% and a specificity of 76%. King & Marrett (1996) found an increased bladder cancer risk with increasing duration of exposure to THMs, but the association was statistically significant and of higher magnitude only after 35 or more years of exposure (Table 32). A statistically significant increased bladder cancer risk was also found for the highest quartile of cumulative exposure to THMs (Table 33). The risk of bladder cancer incidence was about 40% higher among persons exposed to greater than 1956 g of THMs per litre-year in water compared with those exposed to less than 584 g/litre-year. No association between cumulative exposure to THMs and increased bladder cancer risk was found in Colorado (USA) (McGeehin et al., 1993), but the highest cumulative THM levels in Colorado were much lower than those found in Ontario (Table 34). A population-based case-control study conducted in 1986-1989 in Iowa (USA) (Cantor et al., 1998) found associations of increased bladder cancer risk with total and average lifetime exposure to THMs (Table 35), but increased relative risks were restricted to men who had ever smoked (see also section 5.2.1.2). These findings were similar to associations found with duration of exposure to chlorinated surface water. Relative risks for women suggested a protective effect for increased lifetime average THM levels; ORs ranged from 0.6 to 0.9, and 95% CIs ranged from 0.3 to 1.3. Past exposure to chlorination by-products was estimated by combining information about water sources, historical chlorine use, lifetime residential history, fluid consumption, THM levels and other water quality data, including volatile organic compounds, TOC, TOX, pesticides, total dissolved solids and nitrates (Lynch et al., 1990; Neutra & Ostro, 1992). Data reported thus far (Cantor et al., 1998) have been restricted to THMs. Table 32. Bladder cancer risks and estimated maximum annual exposure to trihalomethanes in water in Ontario (Canada)a Years of Odds ratio (95% CI)b exposure THMs >24 g/litre THMs >49 g/litre THMs >74 g/litre 0-9 1.0 1.0 1.0 10-19 1.2 (0.9-1.7) 1.1 (0.9-1.4) 1.1 (0.9-1.4) 20-34 1.2 (0.9-1.5) 1.4 (1.1-1.8) 1.3 (0.9-1.8) >34 1.6 (1.2-2.1) 1.6 (1.1-2.5) 1.7 (1.1-2.7) a From King & Marrett (1996). b Adjusted for age, gender, smoking, education and calorie intake. Table 33. Bladder cancer risks and estimated cumulative exposure to trihalomethanes in Ontario (Canada)a THMs-years of exposure Odds ratio (95% CI)b (g/litre-year) 0-583 1.0 584-1505 1.2 (0.9-1.6) 1506-1956 1.1 (0.8-1.4) 1957-6425 1.4 (1.1-1.9) a From King & Marrett (1996). b Adjusted for age, gender, smoking, education and calorie intake. Table 34. Bladder cancer risks and estimated cumulative exposure to trihalomethanes in Colorado (USA)a THMs-years of exposure Odds ratiob (g/litre-year) 0 1.0 <200 1.8 201-600 1.1 >600 1.8 a From McGeehin et al. (1993). b 95% CI not available; P for trend = 0.16. Table 35. Bladder cancer risks and estimated cumulative exposure to trihalomethanes in Iowa (USA)a Lifetime THM exposure Odds ratio (95% CI)b Total lifetime exposure (g) <0.04 1.0 0.05-0.12 1.3 (1.0-1.6) 0.13-0.34 1.1 (0.9-1.4) 0.35-1.48 1.1 (0.9-1.4) 1.49-2.41 1.2 (0.8-1.7) >2.42 1.3 (0.9-2.0) Lifetime average exposure (g/litre) <0.8 1.0 0.8-2.2 1.2 (1.0-1.5) 2.3-8.0 1.1 (0.8-1.4) 8.1-32.5 1.1 (0.8-1.4) 32.6-46.3 1.3 (0.9-1.8) >46.3 1.2 (0.8-1.8) a From Cantor et al. (1998). b Adjusted for age, gender, study period, smoking, education and high-risk occupation. 2) Colon cancer risk A colon cancer association was studied in Wisconsin (USA) in an interview-based study (Young et al., 1987, 1990) of 347 incident cases of colon cancer, 611 population-based controls and 639 controls with cancer of other sites. White males and females between the ages of 35 and 90 were eligible for selection in the study. Lifetime residential and water source histories and information on water consumption habits, diet, demographic information, medical and occupational histories, lifestyle and other factors was obtained by a self-administered questionnaire and augmented with information from medical records. Historical exposures to THMs were estimated using a predictive statistical model based on current THM levels and water supply operation records. Multivariate logistic regression analysis was used to estimate the risk of colon cancer adjusted for age, gender and urbanization. Individuals exposed to drinking-water containing more than 40 g of THMs per litre, the highest exposure category, were found to be at no greater risk of colon cancer than individuals exposed to water with no or trace levels of THMs. Nor did cumulative exposure to THMs present a colon cancer risk (Table 36). Although this analysis suggests that the presence of THMs in Wisconsin Table 36. Colon cancer risks and estimated cumulative exposure to trihalomethanes in Wisconsin (USA)a Lifetime THM exposure Odds ratiob 95% CI (g/litre-year) <137 1.0 137-410 1.1 0.7-1.8 >410 0.7 0.4-1.2 a From Young et al. (1987). b Adjusted for age, gender and population of place of residence. drinking-water is not associated with a colon cancer risk, THM levels in Wisconsin are low: 98% of water samples had concentrations less than 100 g/litre. Lawrence et al. (1984) studied the relationship of THMs and colo-rectal cancer in New York State (USA) where THM exposure was higher than in Wisconsin and found no association. Included in the study were 395 white female teachers in New York State who died from colo-rectal cancer and an equal number of teachers who died from non-cancer causes. Cumulative chloroform exposure was estimated by a statistical model that considered water treatment operational records during the 20 years prior to death. The distribution of chloroform exposure was not significantly different between cases and controls, and no effect of cumulative exposure was found in a logistic analysis controlling for average source type, population density, marital status, age and year of death. The risk of cancer was not found to be higher among those using a surface water source containing THMs (OR = 1.1; 90% CI = 0.8-1.4). Mean levels of cumulative THM exposure were similar among cases (635 g/litre-years) and controls (623 g/litre-years). The population-based case-control study conducted from 1986 to 1989 in Iowa (USA) (Hildesheim et al., 1998) found no increased risk of colon cancer or any subsites associated with estimates of exposure to THMs, either total (g) or average lifetime (g/litre) (Table 37). A prospective cohort study of 41 836 post-menopausal women in Iowa (USA) (Doyle et al., 1997) included an analysis of women who reported drinking municipal or private well-water for more than the past 10 years ( n = 28 237). Historical water treatment and water quality data were used to ascertain exposure to THMs. The primary source of information on THMs was a 1986-1987 water survey. All women who lived in the same community were assigned the same level of Table 37. Colon cancer rsks and estimated cumulative exposure to trihalomethanes in Iowa (USA)a Lifetime THM exposure Odds ratio (95% CI)b Total lifetime exposure (g) <0.04 1.0 0.05-0.12 1.0 (0.7-1.2) 0.13-0.34 0.9 (0.6-1.2) 0.35-1.48 1.2 (0.9-1.6) 1.49-2.41 0.5 (0.3-0.9) >2.42 1.1 (0.7-1.8) Lifetime average exposure (g/litre) <0.8 1.0 0.8-2.2 1.0 (0.8-1.3) 2.3-8.0 0.9 (0.7-1.3) 8.1-32.5 1.1 (0.8-1.4) 32.6-46.3 0.9 (0.6-1.5) >46.3 1.1 (0.7-1.6) a From Hildesheim et al. (1998). b Adjusted for age and gender. exposure to THMs. A dose-response relationship (Table 38) was found with increasing chloroform levels in municipal drinking-water for all cancers combined, colon cancer, lung cancer and melanoma (test for trend P < 0.05). The highest exposure covered a wide range of values, and selection of the exposure categories for chloroform is unusual and does not reflect normal high exposures to chloroform. Additional analyses after exclusion of women who reported a history of colo-rectal polyps and adjustment for additional risk or protective factors did not change the dose-response relationship and slightly increased the relative risk estimates. No associations were observed between colon cancer and BDCM, DBCM or bromoform, but levels were low, and many water systems had no detectable levels of these THMs. For example, the geometric mean levels of BDCM and DBCM in surface water were 8.7 and 0.4 g/litre, respectively. Because water quality data were also available from a 1979 water survey for a national bladder cancer study, analyses were also conducted using these data. Fewer municipalities were sampled for THMs in 1979, and only 16 461 participants were included. In this analysis, colon cancer was found to be associated with increasing exposure to chloroform levels. Table 38. Risks of cancer incidence associated with chloroform levels in post-menopausal women, Iowa (USA)a Cancer site Chloroform concentration (g/litre) 1-2 3-13 14-287 Cases RRb 95% CI Cases RRb 95% CI Cases RRb 95% CI Bladder 11 0.9 0.4-2.1 12 1.3 0.6-2.8 7 0.7 0.3-1.7 Colon 41 1.1 0.7-1.7 42 1.4 0.9-2.2 57 1.7 1.1-2.6 Rectal 19 0.8 0.4-1.5 14 0.8 0.4-1.5 22 1.1 0.6-2.0 Breast 151 1.1 0.9-1.3 131 1.2 0.9-1.5 136 1.1 0.9-1.4 Kidney 5 0.6 0.2-1.7 9 1.3 0.5-3.2 7 0.9 0.3-2.4 Lung 35 1.4 0.8-2.3 40 2.0 1.2-3.2 42 1.9 1.1-3.0 Melanoma 15 2.5 1.0-6.5 6 1.3 0.4-3.9 17 3.2 1.3-8.2 All cancers 253 1.1 0.9-1.3 220 1.3 1.05-1.5 268 1.3 1.1-1.5 a From Doyle et al. (1997). b Reported relative risks were adjusted for age, education, smoking status, physical activity, fruit and vegetable intake, total energy intake, body mass index and waist to hip ratio. 3) Rectal cancer risk The population-based case-control study conducted in Iowa (USA) (Hildesheim et al., 1998) found an increased risk of rectal cancer associated with estimates of exposure to THMs after controlling for age, gender and average population size for men and women. For total lifetime exposure to THMs of greater than 1.48 g, odds ratios were almost double (OR = 1.9; 95% CI = 1.2-3.0) those for total lifetime exposure to THMs of less than 0.05 g. For lifetime average THM concentrations of 32.6-46.3 g/litre and greater than 46.4 g/litre, odds ratios (OR = 1.7; 95% CI = 1.1-2.6) were almost 70% greater than those for average exposures of less than 0.8 g/litre (Table 39). This is the only study to report increased rectal cancer risks associated with THM exposure. Table 39. Rectal cancer risks and estimated cumulative exposure to trihalomethanesa Lifetime THM exposure Odds ratio (95% CI)b Total lifetime exposure (g) <0.04 1.0 0.05-0.12 1.3 (1.0-1.6) 0.13-0.34 1.3 (0.9-1.8) 0.35-1.48 1.5 (1.1-2.1) 1.49-2.41 1.9 (1.2-3.0) >2.42 1.6 (1.0-2.6) Lifetime average exposure (g/litre) <0.8 1.0 0.8-2.2 1.0 (0.8-1.4) 2.3-8.0 1.2 (0.9-1.7) 8.1-32.5 1.2 (0.9-1.7) 32.6-46.3 1.7 (1.1-2.6) >46.3 1.7 (1.1-2.6) a From Hildesheim et al. (1998). b Adjusted for age and gender. 4) Mutagenicity A cohort study of populations exposed to various levels of mutagenicity in drinking-water and a case-control study of kidney and bladder cancers are currently being conducted in Finland (Tuomisto et al., 1995). No results have yet been reported for the case-control study. The cohort study (Koivusalo et al., 1996, 1997) included 621 431 persons living in the same town in which they were born and having a water connection in 1970. Cancer incidence in the cohort was compared with national cancer incidence stratified by gender, time period and age group. Cases were derived from the population-based Finnish Cancer Registry, and follow-up of the cohort started in 1970. Past exposure to drinking-water mutagenicity and THMs was assessed using historical water quality information. The quantity of mutagenicity was estimated for each 5-year period from 1955 to 1970 using an empirical equation relating mutagenicity and raw water pH, potassium permanganate oxidation and chlorine dose. A good correlation was found between estimated and measured values for the period 1986- 1987 when measures of mutagenicity were available. The quantity of mutagenicity is minor in raw waters and predominantly results from the chlorination process. The Salmonella/microsome assay is used to assess the mutagenicity of the non-volatile, acid/neutral fraction of chlorinated organic material in water. A mutagen of particular concern is MX, a potent mutagen, as measured by strain S. typhimurium TA100, which may be responsible for up to 57% of the mutagenicity in chlorinated drinking-water (Meier et al., 1986, 1987). After adjusting for age, time period, urbanization and social status, an average exposure to mutagenicity in chlorinated water of 3000 net revertants per litre was found to be associated with a statistically significant increased risk in women for cancers of the bladder (RR = 1.5; 95% CI = 1.0-2.2), rectum (RR = 1.4; 95% CI = 1.0-1.9), oesophagus (RR = 1.9; 95% CI = 1.0-3.5) and breast (RR = 1.1; 95% CI = 1.0-1.2). Past exposure to THMs, one group of volatile by-products of chlorination, did not result in statistically significant excess risks (Koivusalo et al., 1996). Although this study found a moderate association between cancers of the bladder, rectum and oesophagus in women and high levels of mutagenicity in drinking-water, the results should be interpreted with caution, as this is the only analytical epidemiological study of water mutagenicity. Significantly increased relative risks were found only for women, and the magnitude of the risks suggests that results may be due to residual uncontrolled confounding. 5.3.1.3 Summary of results of cancer studies Cumulative exposure to THMs was slightly higher in New York (USA) than in Wisconsin (USA), but no increased colon cancer risk associated with THM exposure was observed in either study. Data reported thus far from a study in Iowa (USA) show that colon cancer risk was not associated with estimates of past exposure to THMs, but rectal cancer risk was associated with increasing amounts of lifetime exposure to THMs. Risks were not reported for THM levels found in drinking-water. A cohort study in Iowa found moderately increased risks associated with a wide reported range of chloroform concentrations (14-287 g/litre). In Ontario (Canada), the risk of bladder cancer incidence was about 40% higher among persons exposed to greater than 1956 g of THMs per litre-year in water compared with those exposed to less than 584 g/litre-year. No association between exposure to THM levels and increased bladder cancer risk was found in Colorado (USA). Data reported thus far from a study in Iowa (USA) show that risk of bladder cancer was not associated with estimates of past exposure to chlorination by-products except among men and smokers, where bladder cancer risk increased with duration of exposure after control for cigarette smoking. In Finland, an average exposure to mutagenicity in chlorinated water of 3000 net revertants per litre was found to be associated with an increased risk in women for cancers of the bladder, rectum, oesophagus and breast; however, past exposure to THMs did not result in statistically significant excess cancer risks. THM levels in drinking-water were not reported for Finland, and these results were reported only in an abstract. No increased risk of bladder cancer was associated with THM exposure in Colorado (USA); cumulative exposure to THMs in Colorado was similar to those in New York (USA) and Wisconsin (USA), where no increased risk was found for colon cancer. In Ontario (Canada), THM exposure was much higher, and a moderate increased risk of bladder cancer was found. At this time, the evidence for an association between THM exposure in drinking-water and colon cancer must be considered inconclusive. No evidence is available from epidemiological studies to suggest an increased risk of colon cancer, but studies have been conducted in areas where cumulative exposures were generally low. The evidence for an association between chlorinated water or THM exposure in drinking-water and bladder cancer is limited. No association was found in Colorado, but cumulative exposures were low. In Canada, where cumulative THM exposure was much higher, a moderate increased risk of bladder cancer was found. It is possible that other unmeasured by-products may be associated with bladder cancer risks, as several studies have found an association between chlorinated surface water and bladder cancer but not between THMs and bladder cancer. It is possible that another DBP or a water contaminant other than a DBP is responsible. A plausible alternative explanation for the observed results is that residence in an area served by a chlorinated surface water supply is simply a surrogate for some other unidentified risk factor or characteristic of urban populations that may be associated with an increased risk of cancer. Other potential DBPs may include other volatile organic contaminants, but studies have considered ingestion, not inhalation, exposures, and very few studies have attempted to assess both long-term exposures to chlorinated water and historical water consumption patterns. Accurate long-term exposure assessment is difficult. Exposures to other chlorinated or brominated compounds or HAAs have also not been considered. Additional studies should continue to assess the risk of the non-volatile fraction of organic by-products. Several studies found an elevated bladder cancer risk among persons who consumed more tapwater per day, but increased tapwater consumption appeared to be an independent risk factor for bladder cancer. The association of water and other fluid consumption and bladder cancer also requires additional study. Weak associations reported between cumulative THM exposure and bladder cancer risks do not provide adequate evidence that THMs cause bladder cancer. In two studies, no association was found. A moderate association was reported for rectal cancer and cumulative THMs, but only in a single study. There is no evidence for an association between the other cancer sites studied and THM exposure. 5.3.2 Epidemiological studies of cardiovascular disease and disinfectant by-products A cohort study of 31 000 residents of Washington County, Maryland (USA) found a slightly increased, but not statistically significant, risk (RR = 1.1; 95% CI = 1.0-1.3) of death due to arteriosclerotic heart disease in residents exposed to chlorinated surface water and springs compared with residents of towns where unchlorinated well-water was used. Water sampling during the study found average chloroform levels of 107 g/litre in the Hagerstown water system, but it is not known if these levels accurately represent long-term exposures. Other observational (Zierler et al., 1988; Zeighami et al., 1990a,b; Riley et al., 1995) epidemiological studies of disinfected water evaluated the possible adverse cardiovascular effects of chlorinated or chloraminated drinking-water, but no by-products were measured. 5.3.2.1 Summary of results of cardiovascular studies Epidemiological studies have not evaluated associations between specific DBPs and cardiovascular disease, but there is no evidence of an increased risk caused by chlorinated or chloraminated drinking-water. 5.3.3 Epidemiological studies of adverse reproductive/developmental outcomes and disinfectant by-products In the Savitz et al. (1995) study, dates of pregnancy were used to assign THM levels from the appropriate water supply and for the periods in the pregnancy in which exposures might cause any adverse effect. No associations were reported between THM levels in North Carolina (USA) and estimated dose of THMs with miscarriage, preterm delivery or low birth weight. Savitz et al. (1995) found no increased risk of low infant birth weight or preterm delivery associated with exposure to THM levels of 63-69 g/litre or a computed dose of THMs of 170-1171 g/litre-glasses per day. No increased risk of miscarriage was associated with either (i) THM levels of 60-81.0 or 81.1-168.8 g/litre or (ii) a computed dose of THMs of 140.0-275.0 or 275.1-1171.0 g/litre-glasses per day. Although no increased miscarriage risk was found in this categorical analysis, an analysis using a continuous measure for THMs predicted an association (1.7 per 50 g of THMs per litre increment; 95% CI = 1.1-2.7). This association, however, was not part of an overall dose-response gradient and may be a spurious finding. Another categorical analysis using sextiles of THM exposures showed a much higher miscarriage risk (adjusted OR = 2.8; 95% CI = 1.2-6.1) in the highest sextile but a very low risk or even a possibly protective effect in the next highest sextile (adjusted OR = 0.2; 95% CI = 0.0-0.5). A population-based case-control study in Iowa (USA) used information from birth certificates from January 1989 to June 1990 and a water supply survey conducted in 1987 to study the association of waterborne chloroform and other DBPs with low birth weight (<2500 g), prematurity (<37 weeks' gestation) and intrauterine growth retardation (Kramer et al., 1992). Cases were not mutually exclusive, but each outcome was analysed separately. Controls were randomly selected from the same birth certificates. The study included 159 low birth weight and 795 normal birth weight infants, 342 premature infants and 1710 controls, and 187 intrauterine growth-retarded infants and 935 controls. Mothers were not interviewed to obtain information about their residential history during pregnancy or possible risk factors, and other exposures were not noted on the certificate, which might potentially confound any observed association or modify its effect. Information on maternal age, parity, adequacy of prenatal care, marital status, education and maternal smoking was available from the birth certificate. Residence of the mother at the time of birth determined which water system was used to assign levels of exposure to THM and TOX measured in a previous municipal water survey. The only statistically significant finding was a moderately increased risk (OR = 1.8; 95% CI = 1.1-2.9) of intrauterine growth retardation associated with chloroform levels of greater than 9 g/litre in water after controlling for confounding characteristics from the certificate. Prematurity (OR = 1.1; 95% CI = 0.7-1.6) and low birth weight (OR = 1.3; 95% CI = 0.8-2.2) were not found to be associated with chloroform levels. No statistically significant associations were seen with any of these developmental outcomes and BDCM, DBCM, bromoform or organic halides. Interpretation of the results of this study, however, is limited because the study design was more ecological than analytical. The authors considered the results to be preliminary because of possible bias. The ascertainment and classification of exposure to the water contaminants were imprecise and may have resulted in misclassification bias. Municipal measures of by-products assigned to the residences for exposure purposes may have been either higher or lower than actual exposures. Characteristics that were not identified (e.g., alcohol consumption) could be responsible for confounding bias. A cross-sectional epidemiological study (Bove et al., 1992a, 1995) was conducted in four northern New Jersey (USA) counties to explore possible associations between THM levels and 13 developmental and adverse reproductive outcomes: low birth weight (<2500 g), prematurity (<37 weeks), small for gestational age, very low birth weight (<1500 g), stillbirths, surveillance malformations for 33 selected categories, central nervous system defects and subgroups, oral cleft defects and subgroups, and cardiac defects and subgroups. A total of 143 hypotheses were formally evaluated; as the stated objective of the study was to identify promising leads for further research rather than for decision-analytical purposes, each finding was reported as if it were the sole focus of the study without statistical adjustment for multiple comparisons. Reproductive outcomes over a 4-year period, from January 1985 to December 1988, were obtained from a population-based birth defects registry and vital records, birth certificates and death certificates. A total of 80 938 live births and 594 fetal deaths were studied in 75 towns selected because residents were mostly served by public water systems. All information about reproductive outcomes, potential confounding characteristics and risk factors, such as maternal age, race, education, primipara, previous stillbirth or miscarriage, sex and adequacy of prenatal care, was obtained from vital records. Information on other potential important confounders -- maternal occupation, drug use during pregnancy, smoking and alcohol consumption -- was not available for analysis. Mothers were not interviewed for information on individual exposures and potentially confounding characteristics. Information about water quality for the 75 towns was obtained from existing records. Monthly estimates of each water contaminant in a town's water system were used to assign exposure for each gestational month of each live birth and fetal death. A mother's residence at birth was assumed to be her residence throughout her pregnancy. All water systems in the study were chlorinated, but sufficient groundwater sources were available to ensure inclusion of areas with very low levels of THMs. The drinking-water contaminants studied were THMs, trichloroethylene, tetrachloroethylene, dichloroethylenes, 1,1,1-trichloroethane, carbon tetrachloride, 1,2-dichloroethane, benzene and nitrate. Type of water source was also considered. For evaluation of risk with different levels of exposure, THMs were categorized into six different levels of potential exposure: <20, >20-40, >40-60, >60-80, >80-100 and >100 g/litre. Reported associations with levels of THMs greater than 100 g/litre included small for gestation age (OR = 1.5; 90% CI = 1.2-1.9) and oral cleft defects (OR = 3.2; 90% CI = 1.2-7.3). Reported associations with levels of THMs above 80 g/litre included all surveillance defects (OR = 1.6; 90% CI = 1.2-2.0), central nervous system defects (OR = 2.6; 90% CI = 1.5-4.3), neural tube defects (OR = 3.0; 90% CI = 1.3-6.6) and major cardiac defects (OR = 1.8; 90% CI = 1.0-3.3). Moderate to strong associations were found for central nervous system, oral cleft and neural tube defects, but only a small number of cases were studied. The study included 4082 small for gestational age infants, but fewer numbers with birth defects: 56 infants with neural tube defects, 83 with oral cleft defects, 108 with major cardiac defects and 118 with central nervous system defects. The observed increased risk for other reproductive outcomes was smaller, and these weak associations could be due to unidentified confounding bias. As in the Iowa (USA) study, the ecological study design for assessment of individual exposure and confounding bias limits the interpretation of the results. The investigators noted "the difficulty of interpreting the available water contamination data and the numerous assumptions needed in order to estimate contaminant levels" as a limitation (Bove et al., 1992a). "By itself, this study cannot resolve whether the drinking-water contaminants caused the adverse birth outcomes" (Bove et al., 1995). Also reported were results of a population-based case-control study in New Jersey (USA) to determine risks of cardiac defects, neural tube defects, oral clefts, very low birth weight and low birth weight associated with exposure to different THM levels (Bove et al., 1992b). A total of 563 mothers of cases and controls were interviewed by telephone some 6-54 months after giving birth. The study included 185 infants with birth defects, 37 of whom had neural tube defects, 97 infants with very low birth weights, 113 infants with low birth weights and 138 infants of normal weight without birth defects. Information was obtained for a potential exposure period 3 months prior to conception through the end of pregnancy and included residences of the mother, sources of drinking-water, tapwater consumption, showering and smoking habits, alcohol consumption, exposures in and around the home, prescription drugs, medical history and previous adverse reproductive outcomes. For neural tube defects, a 4-fold increased risk was found to be associated with THM levels greater than 80 g/litre (OR = 4.25; 95% CI = 1.0-17.7); however, this estimate is based on only 7 cases and 14 controls. The sample size for the case-control study was small, resulting in low statistical power and a lack of precision in estimating the risk; most importantly, a majority (53.3%) of the mothers of the cases and controls could not be located for interviews, causing possible selection bias. Once contact was established with the mother of the case or control, only 78% agreed to be interviewed. In addition, because of the long period between pregnancy and the interview, inaccurate recall of the mothers may be responsible for incorrect or biased information about both potential exposures and confounding characteristics. An assessment of selection bias by the investigators showed that risk had likely been overestimated for neural tube defects. After correcting for selection bias, the risk associated with THM levels greater than 80 g/litre was reduced from a 4-fold to a 50% increased risk. The observed risk now represents a weak association where unknown confounding bias might be responsible. A second population-based case-control study was recently reported in a published abstract (Klotz et al., 1996). Cases were births or fetal deaths after 20 weeks' gestation in 1993 and 1994 with anencephaly, spina bifida or encephalocele reported to the New Jersey Birth Defects Registry or Vital Statistics. Controls were randomly selected by month of birth from birth certificates during the same period as cases. In home interviews, information was obtained on ingestion and non-ingestion exposures, other environmental and occupational exposures, and pregnancy characteristics. Exposures were estimated from water quality monitoring data for the appropriate public water system selected to approximate the critical time for neural closure (fourth week of gestation), similar water quality data from the same source 1 year later, and analysis of residential tapwater collected about 1 year after the critical time. To prevent misclassification of exposure, biological monitoring of urine and exhaled breath from a sample of participants was also conducted. Increased relative risks for neural tube defects were associated with THMs; ORs were generally between 1.5 and 2.1 (Klotz & Pyrch, 1998). The only statistically significant results were observed in infants born with neural tube defects (and no other malformations) and whose mothers' residence in early pregnancy was in an area where the THM levels were greater than 40 g/litre (OR = 2.1; 95% CI = 1.1-4.0). No association was observed between HAAs, HANs or nitrates in drinking-water and risk of neural tube defects. A cohort study (Waller et al., 1998; Swan, 1998) of women members of Kaiser Permanente Medical Care Program in California (USA) evaluated associations of THMs and spontaneous abortion, low birth weight, preterm delivery and intrauterine growth retardation. Results of the spontaneous abortion analysis were recently reported. Information about water consumption, water source and THM levels was collected for the participating cohort members. Data for THM levels in tapwater were obtained from each public water supply for a 3-year period when distribution taps were monitored at least quarterly. A woman's exposure to THMs was estimated using the average level of THMs for each water supply of samples collected within the woman's first trimester (77% of the cohort) or within 30 days of the subject's first trimester (4%) or the annual average from the utility's annual water quality report (9%). High first-trimester levels of THMs were based on levels corresponding to the 75th percentile, i.e., >75 g of total THMs per litre, >16 g of chloroform per litre, >15 g of bromoform per litre, >17 g of BDCM per litre and >31 g of DBCM per litre. Pregnancy outcomes were obtained from hospital records, the California Birth Registry and follow-up interviews. Interviews were also used to obtain information about possible confounding characteristics. The study found that an increased risk of miscarriage was associated with a high consumption of water (five or more glasses of cold tapwater per day) containing high levels of total THMs (>75 g/litre), especially for waters high in BDCM (>17 g/litre). In the women with high exposures to total THMs (THM levels >75 g/litre and five or more glasses of water consumed per day), the relative risk of a miscarriage was almost twice that of the low-exposure groups (total THM levels below 75 g/litre and fewer than five glasses of water per day). Of the four THMs, only high exposure to BDCM (>17 g/litre and five or more glasses of water per day) was associated with a 3-fold increased relative risk of miscarriage. The risk for unemployed women was also found to be greater than that for employed women, suggesting that women with a greater opportunity to consume home tapwater with a high level of THMs may be at greater risk. However, it is not certain that other characteristics of employed women were adequately assessed and controlled for (e.g., a healthy worker effect). Preliminary analyses were available for a cross-sectional study (Nuckols et al., 1995) of births to women in Northglenn, Colorado (USA) exposed to chlorinated water from Stanley Lake with high THM levels (32-72 g/litre) and women in Westminster exposed to chloraminated water from Stanley Lake with low THM levels (<20 g/litre). Gallagher et al. (1997)1 used a geographic information system and water quality modelling as described by Nuckols et al. (1995) to conduct a retrospective cohort study in the same water districts. Information about developmental outcomes and possible confounding characteristics was obtained from birth certificates, and exposures to THMs were modelled based on hydraulic characteristics of the water system and THM levels obtained from a monitoring program. However, sufficient information was available to estimate THM exposures for mothers in only 28 census blocks, 11 of 26 census blocks from Northglenn and 17 of 60 blocks from Westminster. The exclusion of such a large number of births from the study seriously limits the interpretation of the observed results. Women with high or low THM exposure may have been selectively excluded, but this could not be determined with the information reported. After excluding births to mothers in blocks with no THM data and births less than 400 g, only 25% of births remained for analysis, and the reported epidemiological associations must be considered inconclusive. 5.3.3.1 Summary of results of reproductive/developmental studies No associations were reported in North Carolina (USA) for THM levels and estimated dose of THMs, but associations between chloroform or THM levels in water and adverse reproductive or developmental outcomes were reported in studies conducted in California, Iowa and New Jersey (USA). A recently completed case-control study (but not yet reported in the peer review literature) in New Jersey (USA) reported increased relative risks for infants born with neural tube defects (and no other malformations) and whose mothers' residence in early pregnancy was in an area where the THM levels were greater than 40 g/litre (OR = 2.1; 95% CI = 1.1-4.0). Replication of these results in another geographic area is required before causality can be assessed. Previous epidemiological studies in New Jersey reported increased relative risks of neural tube defects associated with the mother's residence in areas with high THMs; however, these studies suffer from methodological limitations, and their results are inconclusive. 1 Gallagher MD, Nuckols JR, Stallones L, & Savitz DA (1997) Exposure to trihalomethanes and adverse pregnancy outcomes in Colorado (Unpublished manuscript). The Waller et al. (1998) study is well designed and well conducted; as it is the first study to suggest an adverse reproductive effect associated with a brominated by-product, results should be used to support further research on spontaneous abortion risks and drinking-water contaminants, including THM species, other DBPs and other water contaminants. The authors discussed the weaknesses and strengths of their study. One weakness is that water consumption was not assessed at work, but investigators did consider risks separately for women who worked outside the home, reporting that "results were stronger in women not employed outside the home, for whom our home based exposure assessment should be more precise" (Waller et al., 1998). A major limitation is that DBPs other than THMs and other water contaminants were not studied. A concern is how to interpret the results of the Waller et al. (1998) study in light of the findings of Swan et al. (1998). Swan et al. (1998) analysed tapwater and bottled water consumption in the same cohort studied by Waller et al. (1998) and reported a dose-related increase in spontaneous abortions among tapwater drinkers in Region I, but not in Region II or III. "Our prior studies suggested that the relation between spontaneous abortion and tapwater was independent of chlorination by-products, since the strongest associations were seen in the two studies conducted in areas served only by unchlorinated groundwater. Additionally, in the two rodent studies we conducted, a trend toward increased rates of fetal resorption was seen in rats drinking unchlorinated groundwater, compared with bottled water. In our current study, as discussed in the study by Waller et al., spontaneous abortion risk was increased by exposure to specific chlorination by-products in all regions. Nevertheless, we believe that the associations with cold tapwater and bottled water presented here, which are specific to Region I, cannot be explained by exposure to chlorination by-products, because the association is seen in the absence of high levels of these chemicals" (Swan et al., 1998). In a letter to the editor, Swan & Waller (1998) suggested that there may be "some constituent in addition to THMs" that is specific to tapwater in Region I but provided a poor explanation of why this may be so. "Bromodichloromethane (or some compound highly correlated with it) was the trihalomethane most strongly associated with SAB [spontaneous abortion]" (Waller et al., 1998). "Swan et al. found a dose-related increase in SABs among tapwater drinkers in Region I, but not in Regions II or III. Exposure to TTHM [total THMs] or bromodichloromethane does not entirely explain this association, since a tapwater effect is still evident among Region I women with low levels of both TTHM and bromodichloromethane. Furthermore, the initial studies in Region I found the strongest effect in areas served only by unchlorinated groundwater. Thus, it is likely that other factors contributed to the tapwater effect described by Swan et al." (Waller et al., 1998). Because of these concerns, judgement about the interpretation of the results of Waller et al. (1998) should be deferred until additional water quality data are analysed for the cohort or another well designed and well conducted epidemiological study finds similar results. Although the study by Waller et al. (1998) does not provide sufficient evidence for a cause-effect relationship between exposure to THMs and early-term miscarriages, it does provide important new information that should be pursued with additional research. Assessing the causality of an observed epidemiological association requires evidence from more than a single study, and additional information is needed on other water exposures. Several exploratory epidemiological studies have suggested that certain adverse reproductive effects and developmental outcomes may be associated with chloroform or THMs in drinking-water, but additional studies are required to determine whether these observed associations are spurious or due to possible bias. The latest study from New Jersey (USA) reported a moderate association between neural tube defects and THM levels of more than 40 g/litre but no associations with HAAs, HANs or nitrates in drinking-water. The Waller et al. (1998) and Klotz & Pyrch (1998) studies require replication in another area before results can be properly interpreted. 5.4 Summary Epidemiological studies have not identified an increased risk of cardiovascular disease associated with chlorinated or chloraminated drinking-water. Based on the entire cancer-chlorinated drinking-water epidemiology database, there is better evidence for an association between exposure to chlorinated surface water and bladder cancer than for other types of cancer. However, the latest published study (Cantor et al., 1998) notes several inconsistencies in results among the studies for smokers/non-smokers and males/females, and the evidence is still considered insufficient to allow a judgement as to whether the association is causal and which water contaminants may be important. Evidence for a role of THMs is weak. Poole (1997) also notes that "The basic conclusion of the present report is that the hypothesis of a causal relationship between consumption of chlorination by-products and the risk of any cancer, including bladder cancer and rectal cancer, is still an open question." The overall findings of Cantor et al. (1998) support the hypothesis of an association between bladder cancer and duration of use of chlorinated surface water or groundwater and estimated THM exposures, but aspects of these results caution against a simple interpretation and raise additional questions about the nature of the association. An increase in bladder cancer risk was found with duration of chlorinated groundwater use, as well as with total duration of chlorinated drinking-water (surface water plus groundwater) use, with relative risks similar to those observed with chlorinated surface water. This finding is unexpected, because the levels of by-products from most chlorinated groundwaters are much lower than those in treated surface water. In addition, risk was found to increase with duration of chlorinated surface water use among ever-smokers, but not never-smokers, and among men, but not women. This raises questions of internal consistency, as well as consistency with other findings. In contrast, Cantor et al. (1998) found associations for both sexes, primarily among never-smokers. Cantor et al. (1985) noted: "In Ontario, King and Marrett noted somewhat higher risk estimates for never-smokers associated with duration of chlorinated surface water. In Colorado, McGeehin et al. reported similar patterns of risk among smokers and never-smokers, and among men and women. Finally, in a case-control study from Washington County, Maryland, Freedman et al. reported results that parallel the current findings, namely that the risk associated with chlorinated surface water was primarily observed among men and among smokers. Reasons for differences among these observations and differences with results from our study are unclear. A possible explanation for the apparent discrepancies in findings for smokers and never-smokers among studies may reside in water quality and water treatment differences in the respective study areas, with resulting variations in the chemical composition of byproduct mixtures. Nevertheless, results should not differ by sex." The existing epidemiological data are insufficient to allow a conclusion that the observed associations between bladder or any other cancer and chlorinated drinking-water or THMs are causal or provide an accurate estimate of the magnitude of risk. Any association between exposure to chlorinated surface water, THMs or the mutagenicity of drinking-water and cancer of the colon, rectum, pancreas, brain and other sites cannot be evaluated at this time because of inadequate epidemiological evidence. However, the findings from well conducted studies associating bladder cancer with chlorinated water and THMs cannot be completely dismissed, even though inconsistencies have been noted for risks among men and women and among smokers and non-smokers. Because of the large number of people exposed to chlorinated drinking-water, it is important to resolve this issue using studies designed with sound epidemiological principles. Additional studies to resolve the questions about the associations that have been reported for chlorinated surface water, THMs, fluid and tapwater consumption and bladder cancer and reproductive and developmental effects must focus on the resolution of various problems noted in previous studies, especially consideration of exposures to other DBPs. The existing epidemiological data are insufficient to allow the importance of the observed associations of chlorinated drinking-water or THMs and adverse pregnancy outcomes to be assessed. Although several studies have suggested that increased risks of neural tube defects and miscarriage may be associated with THMs or selected THM species, additional studies are needed to determine whether the observed associations are spurious. A recently convened scientific panel (US EPA, 1997) concluded that the results of published epidemiological studies do not provide convincing evidence that DBPs cause adverse pregnancy outcomes. The panel recommended that additional studies be conducted, specifically that the Waller et al. (1998) study be expanded to include additional exposure information about by-products other than THMs and that a similar study be conducted in another geographic area. 6. RISK CHARACTERIZATION It should be noted that the use of chemical disinfectants in water treatment usually results in the formation of chemical by-products, some of which are potentially hazardous. However, the risks to health from these by-products at the levels at which they occur in drinking-water are extremely small in comparison with the risks associated with inadequate disinfection. Thus, it is important that disinfection not be compromised in attempting to control such by-products. 6.1 Characterization of hazard and dose-response 6.1.1 Toxicological studies 6.1.1.1 Chlorine A WHO Working Group for the Guidelines for drinking-water quality (WHO, 1993) considered chlorine. This Working Group determined a tolerable daily intake (TDI) of 150 g/kg of body weight for free chlorine. This TDI is derived from a NOAEL of approximately 15 mg/kg of body weight per day in 2-year studies in rats and mice (NTP, 1992), incorporating an uncertainty factor of 100 (10 each for intra- and interspecies variation). There are no new data that indicate that this TDI should be changed. 6.1.1.2 Monochloramine A WHO Working Group for the Guidelines for drinking-water quality considered monochloramine (WHO, 1993). This Working Group determined a TDI of 94 g/kg of body weight based on a NOAEL of approximately 9.4 mg/kg of body weight per day, the highest dose tested, in a 2-year bioassay in rats (NTP, 1992), incorporating an uncertainty factor of 100 (10 each for intra- and interspecies variation). There are no new data that indicate that this TDI should be changed. 6.1.1.3 Chlorine dioxide Chlorine dioxide chemistry in drinking-water is complex, but the major breakdown product in drinking-water is chlorite. In establishing a specific TDI for chlorine dioxide, data on both chlorine dioxide and chlorite can be considered, given the rapid hydrolysis to chlorite. Therefore, an oral TDI for chlorine dioxide is 30 g/kg of body weight, based on the NOAEL of 2.9 mg/kg of body weight per day for neurodevelopmental effects of chlorite in rats (CMA, 1997). 6.1.1.4 Trihalomethanes Cancer following chronic exposure is the primary hazard of concern for this class of DBPs. Owing to the weight of evidence indicating that chloroform can induce cancer in animals only after chronic exposure to cytotoxic doses, it is clear that exposures to low concentrations of chloroform in drinking-water do not pose carcinogenic risks. The NOAEL for cytolethality and regenerative hyperplasia in mice was 10 mg/kg of body weight per day after administration of chloroform in corn oil for 3 weeks (Larson et al., 1994b). Based on the mode of action evidence for chloroform carcinogenicity, a TDI of 10 g/kg of body weight was derived using the NOAEL for cytotoxicity in mice and applying an uncertainty factor of 1000 (10 each for inter- and intraspecies variation and 10 for the short duration of the study). This approach is supported by a number of additional studies. This TDI is similar to the TDI derived in the Guidelines for drinking-water quality (WHO, 1998), which was based on a 7.5-year study in dogs. In this study, beagle dogs were given chloroform in a toothpaste base in gelatin capsules, 6 days per week for 7.5 years, at 0, 15 or 30 mg/kg of body weight per day. Slight hepatotoxicity was observed at 15 mg/kg of body weight per day (Heywood et al., 1979). Incorporating an uncertainty factor of 1000 (10 each for intra- and interspecies variation and 10 for use of a LOAEL rather than a NOAEL and a subchronic study), a TDI of 13 g/kg of body weight (corrected for 6 days per week dosing) was derived. Among the brominated THMs, BDCM is of particular interest because it produces tumours in rats and mice and at several sites (liver, kidney and large intestine) after corn oil gavage (NTP, 1987). The induction of colon tumours in rats by BDCM (and by bromoform) is also interesting because of the epidemiological associations with colo-rectal cancer (see section 5.3.1). BDCM and the other brominated THMs are also weak mutagens (IARC, 1991, 1999; Pegram et al., 1997). It is generally assumed that mutagenic carcinogens will produce linear dose-response relationships at low dose, as mutagenesis is generally considered to be an irreversible and cumulative effect. In a 2-year bioassay, BDCM given by corn oil gavage induced tumours (in conjunction with cytotoxicity and increased proliferation) in the kidneys of mice and rats at doses of 50 and 100 mg/kg of body weight per day, respectively (NTP, 1987). The large intestine tumours in rats occurred after exposure to both 50 and 100 mg/kg of body weight per day. Using the incidence of kidney tumours in male mice from this study, quantitative risk estimates have been calculated, yielding a slope factor1 of 4.8 10-3 [mg/kg of body weight per day]-1 and a calculated dose of 2.1 g/kg of body weight per day for a risk level of 10-5 (IRIS, 1993). A slope factor of 4.2 10-3 [mg/kg of body weight per day]-1 (2.4 g/kg of body weight per day for a 10-5 risk) was derived based on the incidence of large intestine carcinomas in the male rat. IARC (1991, 1999) has classified BDCM in Group 2B (possibly carcinogenic to humans). 1 Slope factors given here do not incorporate a surface area to body weight correction. DBCM and bromoform were studied in long-term bioassays. In a 2-year corn oil gavage study, DBCM induced hepatic tumours in female mice, but not in rats, at a dose of 100 mg/kg of body weight per day (NTP, 1985). In previous evaluations, it has been suggested that the corn oil vehicle may play a role in the induction of tumours in female mice (WHO, 1996). A small increase in tumours of the large intestine in rats was observed in the bromoform study at a dose of 200 mg/kg of body weight per day. No neoplastic effects were associated with exposure of mice to chloroform (NTP, 1989a). The slope factors based on these tumours are 6.5 10-3 [mg/kg of body weight per day]-1 for DBCM or 1.5 g/kg of body weight per day for 10-5 risk (IRIS, 1992) and 1.3 10-3 [mg/kg of body weight per day]-1 or 7.7 g/kg of body weight per day for 10-5 risk for bromoform (IRIS, 1991). These two brominated THMs are weakly mutagenic in a number of assays, and they were by far the most mutagenic DBPs of the class in the GST-mediated assay system (DeMarini et al., 1997; Pegram et al., 1997). Because they are the most lipophilic THMs, additional concerns about whether corn oil may have affected their bioavailability in the long-term studies should be considered. A NOAEL for DBCM of 30 mg/kg of body weight per day has been established in a 13-week corn oil gavage study, based on the absence of histopathological effects in the liver of rats (NTP, 1985). A TDI for DBCM of 30 g/kg of body weight was derived based on the NOAEL for liver toxicity of 30 mg/kg of body weight per day and an uncertainty factor of 1000 (10 each for inter- and intraspecies variation and 10 for the short duration of the study and possible carcinogenicity). IARC (1991, 1999) has classified DBCM in Group 3 (not classifiable as to its carcinogenicity to humans). A NOAEL for bromoform of 25 mg/kg of body weight per day can be derived on the basis of the absence of liver lesions in rats after 13 weeks of dosing by corn oil gavage (NTP, 1989a). A TDI for bromoform of 25 g/kg of body weight was derived based on this NOAEL for liver toxicity and an uncertainty factor of 1000 (10 each for inter- and intraspecies variation and 10 for the short duration of the study and possible carcinogenicity). IARC (1991, 1999) has classified bromoform in Group 3 (not classifiable as to its carcinogenicity to humans). 6.1.1.5 Haloacetic acids The induction of mutations by DCA is very improbable at the low doses that would be encountered in chlorinated drinking-water. The available data indicate that DCA differentially affects the replication rates of normal hepatocytes and hepatocytes that have been initiated (Pereira & Phelps, 1996). The dose-response relationships are complex, with DCA initially stimulating division of normal hepatocytes. However, at the lower chronic doses used in animal studies (but still very high relative to those that would be derived from drinking-water), the replication rate of normal hepatocytes is eventually sharply inhibited. This indicates that normal hepatocytes eventually down-regulate those pathways that are sensitive to stimulation by DCA. However, the effects in altered cells, particularly those that express high amounts of a protein that is immunoreactive to a c-Jun antibody, do not seem to be able to down-regulate this response (Stauber & Bull, 1997). Thus, the rates of replication in the pre-neoplastic lesions with this phenotype are very high at the doses that cause DCA tumours to develop with a very low latency. Preliminary data suggest that this continued alteration in cell birth and death rates is also necessary for the tumours to progress to malignancy (Bull et al., 1990). This interpretation is supported by studies that employ initiation/promotion designs as well (Pereira, 1996). Based upon the above considerations, it is suggested that the currently available cancer risk estimates for DCA be modified by incorporation of newly developing information on its comparative metabolism and modes of action to formulate a biologically based dose-response model. These data are not available at this time, but they should become available within the next 2-3 years. The effects of DCA appear to be closely associated with doses that induce hepatomegaly and glycogen accumulation in mice. The LOAEL for these effects in an 8-week study in mice was 0.5 g/litre, corresponding to approximately 100 mg/kg of body weight per day, and the NOAEL was 0.2 g/litre, or approximately 40 mg/kg of body weight per day (Kato-Weinstein et al., 1998). A TDI of 40 g/kg of body weight has been calculated by applying an uncertainty factor of 1000 to this NOAEL (10 each for inter- and intraspecies variation and 10 for the short duration of the study and possible carcinogenicity). IARC (1995) has classified DCA in Group 3 (not classifiable as to its carcinogenicity to humans). TCA is one of the weakest activators of the PPAR known (Issemann & Green, 1990). It appears to be only marginally active as a peroxisome proliferator, even in rats (DeAngelo et al., 1989). Furthermore, treatment of rats with high levels of TCA in drinking-water does not induce liver tumours (DeAngelo et al., 1997). These data strongly suggest that TCA presents little carcinogenic hazard to humans at the low concentrations found in drinking-water. From a broader toxicological perspective, the developmental effects of TCA are the end-point of concern (Smith et al., 1989a; Saillenfait et al., 1995). Animals appear to tolerate concentrations of TCA in drinking-water of 0.5 g/litre (approximately 50 mg/kg of body weight per day) with little or no signs of adverse effect. At 2 g/litre, the only sign of adverse effect appears to be hepatomegaly. The hepatomegaly is not observed in mice at doses of 0.35 g of TCA per litre in drinking-water, estimated to be equivalent to 40 mg/kg of body weight per day (Pereira, 1996). In a study by Smith et al. (1989a), soft tissue anomalies were observed at approximately 3 times the rate in controls at the lowest dose administered of 330 mg/kg of body weight per day. At this dose, the anomalies were mild and would clearly be in the range where hepatomegaly (and carcinogenic effects) would occur. Considering the fact that the PPAR interacts with cell signalling mechanisms that can affect normal developmental processes, a common mechanism underlying hepatomegaly and the carcinogenic and developmental effects of this compound should be considered. The TDI for TCA is based on a NOAEL estimated to be 40 mg/kg of body weight per day for hepatic toxicity in a long-term study in mice (Pereira, 1996). Application of an uncertainty factor of 1000 to the estimated NOAEL (10 each for inter- and intraspecies variation and 10 for possible carcinogenicity) gives a TDI of 40 g/kg of body weight. IARC (1995) has classified TCA in Group 3 (not classifiable as to its carcinogenicity to humans). Data on the carcinogenicity of brominated acetic acids are too preliminary to be useful in risk characterization. Data available in abstract form suggest, however, that the doses required to induce hepatocarcinogenic responses in mice are not dissimilar to those of the chlorinated acetic acids (Bull & DeAngelo, 1995). In addition to the mechanisms involved in DCA- and TCA-induced cancer, it is possible that increased oxidative stress secondary to their metabolism might contribute to their effects (Austin et al., 1996; Parrish et al., 1996). There are a significant number of data on the effects of DBA on male reproduction. No effects were observed in rats at doses of 2 mg/kg of body weight per day for 79 days, whereas an increased retention of step 19 spermatids was observed at 10 mg/kg of body weight per day. Higher doses led to progressively more severe effects, including marked atrophy of the seminiferous tubules at 250 mg/kg of body weight per day, which was not reversed 6 months after treatment was suspended (Linder et al., 1997b). A TDI of 20 g/kg of body weight was determined by allocating an uncertainty factor of 100 (10 each for inter- and intraspecies variation) to the NOAEL of 2 mg/kg of body weight per day. 6.1.1.6 Chloral hydrate In a 2-year study, chloral hydrate at 1 g/litre of drinking-water (166 mg/kg of body weight per day) induced liver tumours in male mice (Daniel et al., 1992a). Lower doses have not been evaluated. Chloral hydrate has been shown to induce chromosomal anomalies in several in vitro tests but has been largely negative when evaluated in vivo (IARC, 1995). It is probable that the liver tumours induced by chloral hydrate involve its metabolism to TCA and/or DCA. As discussed previously, these compounds are considered to act as tumour promoters. IARC (1995) has classified chloral hydrate in Group 3 (not classifiable as to its carcinogenicity to humans). Chloral hydrate administered to rats for 90 days in drinking-water induced hepatocellular necrosis at concentrations of 1200 mg/litre and above, with no effect being observed at 600 mg/litre (approximately 60 mg/kg of body weight per day) (Daniel et al., 1992b). Hepatomegaly was observed in male mice at doses of 144 mg/kg of body weight per day administered by gavage for 14 days. No effect was observed at 14.4 mg/kg of body weight per day in the 14-day study, but mild hepatomegaly was observed when chloral hydrate was administered in drinking-water at 70 mg/litre (16 mg/kg of body weight per day) in a 90-day follow-up study (Sanders et al., 1982). An uncertainty factor of 1000 (10 each for inter- and intraspecies variation and 10 for the use of a LOAEL instead of a NOAEL) applied to this value gives a TDI of 16 g/kg of body weight. 6.1.1.7 Haloacetonitriles Without appropriate human data or an animal study that involves a substantial portion of an experimental animal's lifetime, there is no generally accepted basis for estimating carcinogenic risk from the HANs. Data developed in subchronic studies provided some indication of NOAELs for the general toxicity of DCAN and DBAN. NOAELs of 8 and 23 mg/kg of body weight per day were identified in 90-day studies in rats for DCAN and DBAN, respectively, based on decreased body weights at the next higher doses of 33 and 45 mg/kg of body weight per day, respectively (Hayes et al., 1986). A Working Group for the WHO Guidelines for drinking-water quality considered DCAN and DBAN (WHO, 1993). This Working Group determined a TDI of 15 g/kg of body weight for DCAN based on a NOAEL of 15 mg/kg of body weight per day in a reproductive toxicity study in rats (Smith et al., 1989b) and incorporating an uncertainty factor of 1000 (10 each for intra- and interspecies variation and 10 for the severity of effects). Reproductive and developmental effects were observed with DBAN only at doses that exceeded those established for general toxicity (about 45 mg/kg of body weight per day) (Smith et al., 1987). A TDI of 23 g/kg of body weight was calculated for DBAN based on the NOAEL of 23 mg/kg of body weight per day in the 90-day study in rats (Hayes et al., 1986) and incorporating an uncertainty factor of 1000 (10 each for intra- and interspecies variation and 10 for the short duration of the study). There are no new data that indicate that these TDIs should be changed. LOAELs for TCAN were identified at 7.5 mg/kg of body weight per day for embryotoxicity and 15 mg/kg of body weight per day for developmental effects in rats (Smith et al., 1988). However, later studies suggest that these responses were dependent upon the vehicle used (Christ et al., 1996). No TDI can be established for TCAN. There are no data useful for risk characterization purposes for other members of the HANs. 6.1.1.8 MX The mutagen MX has recently been studied in a long-term study in rats in which some carcinogenic responses were observed (Komulainen et al., 1997). These data indicate that MX induces thyroid and bile duct tumours. An increased incidence of thyroid tumours was seen at the lowest dose of MX administered (0.4 mg/kg of body weight per day). The induction of thyroid tumours with high-dose chemicals has long been associated with halogenated compounds. The induction of thyroid follicular tumours could involve modifications in thyroid function or mutagenic mode of action. A dose-related increase in the incidence of cholangiomas and cholangiocarcinomas was also observed, beginning at the low dose in female rats, with a more modest response in male rats. The increase in cholangiomas and cholangiocarcinomas in female rats was utilized to derive a slope factor for cancer. The 95% upper confidence limit for a 10-5 lifetime risk based on the linearized multistage model was calculated to be 0.06 g/kg of body weight per day. 6.1.1.9 Chlorite The primary and most consistent finding arising from exposure to chlorite is oxidative stress resulting in changes in the red blood cells (Heffernan et al., 1979a; Harrington et al., 1995a). This end-point is seen in laboratory animals and, by analogy with chlorate, in humans exposed to high doses in poisoning incidents. There are sufficient data available to estimate a TDI for humans exposed to chlorite, including chronic toxicity studies and a two-generation reproductive toxicity study. Studies in human volunteers for up to 12 weeks did not identify any effect on blood parameters at the highest dose tested, 36 g/kg of body weight per day (Lubbers & Bianchine, 1984; Lubbers et al., 1984a). Because these studies do not identify an effect level, they are not informative for establishing a margin of safety. In a two-generation study in rats, a NOAEL of 2.9 mg/kg of body weight per day was identified based on lower auditory startle amplitude, decreased absolute brain weight in the F1 and F2 generations and altered liver weights in two generations (CMA, 1997). Application of an uncertainty factor of 100 to this NOAEL (10 each for inter- and intraspecies variation) gives a TDI of 30 g/kg of body weight. This TDI is supported by the human volunteer studies. 6.1.1.10 Chlorate Like chlorite, the primary concern with chlorate is oxidative damage to red blood cells. Also like chlorite, 0.036 mg/kg of body weight per day of chlorate for 12 weeks did not result in any adverse effect in human volunteers (Lubbers et al., 1981). Although the database for chlorate is less extensive than that for chlorite, a recent well conducted 90-day study in rats is available, which identified a NOAEL of 30 mg/kg of body weight per day based on thyroid gland colloid depletion at the next higher dose of 100 mg/kg of body weight per day (McCauley et al., 1995). A TDI is not derived because a long-term study is in progress, which should provide more information on chronic exposure to chlorate. 6.1.1.11 Bromate Bromate is an active oxidant in biological systems and has been shown to cause an increase in renal tumours, peritoneal mesotheliomas and thyroid follicular cell tumours in rats and, to a lesser extent, hamsters, and only a small increase in kidney tumours in mice. The lowest dose at which an increased incidence of renal tumours was observed in rats was 6 mg/kg of body weight per day (DeAngelo et al., 1998). Bromate has also been shown to give positive results for chromosomal aberrations in mammalian cells in vitro and in vivo but not in bacterial assays for point mutation. An increasing body of evidence, supported by the genotoxicity data, suggests that bromate acts by generating oxygen radicals in the cell. In the WHO Guidelines for drinking-water quality, the linearized multistage model was applied to the incidence of renal tumours in a 2-year carcinogenicity study in rats (Kurokowa et al., 1986a), although it was noted that if the mechanism of tumour induction is oxidative damage in the kidney, application of the low-dose cancer model may not be appropriate. The calculated upper 95% confidence interval for a 10-5 risk was 0.1 g/kg of body weight per day (WHO, 1993). The no-effect level for the formation of renal cell tumours in rats is 1.3 mg/kg of body weight per day (Kurokowa et al., 1986a). If this is used as a point of departure from linearity and if an uncertainty factor of 1000 (10 each for inter- and intraspecies variation and 10 for possible carcinogenicity) is applied, a TDI of 1 g/kg of body weight can be calculated. This compares with the value of 0.1 g/kg of body weight per day associated with an excess lifetime cancer risk of 10-5. At present, there are insufficient data to allow a decision as to whether bromate-induced tumours are a result of cytotoxicity and reparative hyperplasia or a genotoxic effect. IARC (1986, 1987) has assigned potassium bromate to Group 2B: the agent is possibly carcinogenic to humans. 6.1.2 Epidemiological studies Epidemiological studies must be carefully evaluated to ensure that observed associations are not due to bias and that the design is appropriate for an assessment of a possible causal relationship. Causality can be evaluated when there is sufficient evidence from several well designed and well conducted studies in different geographic areas. Supporting toxicological and pharmacological data are also important. It is especially difficult to interpret epidemiological data from ecological studies of disinfected drinking-water, and these results are used primarily to help develop hypotheses for further study. Results of analytical epidemiological studies are insufficient to support a causal relationship for any of the observed associations. It is especially difficult to interpret the results of currently published analytical studies because of incomplete information about exposures to specific water contaminants that might confound or modify the risk. Because inadequate attention has been paid to assessing water contaminant exposures in epidemiological studies, it is not possible to properly evaluate increased relative risks that were reported. Risks may be due to other water contaminants or to other factors for which chlorinated drinking-water or THMs may serve as a surrogate. 6.2 Characterization of exposure 6.2.1 Occurrence of disinfectants and disinfectant by-products Disinfectant doses of several milligrams per litre are typically employed, corresponding to doses necessary to inactivate microorganisms (primary disinfection) or to maintain a distribution system residual (secondary disinfection). A necessary ingredient for an exposure assessment is DBP occurrence data. Unfortunately, there are few published international studies that go beyond case-study or regional data. Occurrence data suggest, on average, an exposure in chlorinated drinking-water to total THMs of about 35-50 g/litre, with chloroform and BDCM being the first and second most dominant species. Exposure to total HAAs can be approximated by a total HAA concentration (sum of five species) corresponding to about one-half of the total THM concentration (although this ratio can vary significantly); DCA and TCA are the first and second most dominant species. In waters with a high bromide to TOC ratio and/or a high bromide to chlorine ratio, greater formation of brominated THMs and HAAs can be expected. When a hypochlorite solution (versus chlorine gas) is used, chlorate may also occur in the hypochlorite solution and be found in chlorinated water. DBP exposure in chloraminated water is a function of the mode of chloramination, with the sequence of chlorine followed by ammonia leading to the formation of (lower levels of) chlorine DBPs (i.e., THMs and HAAs) during the free-chlorine period; however, the suppression of chloroform and TCA formation is not paralleled by a proportional reduction in DCA formation. All factors being equal, bromide concentration and ozone dose are the best predictors of bromate formation during ozonation, with about a 50% conversion of bromide to bromate. A study of different European water utilities showed bromate levels in water leaving operating water treatment plants of less than the detection limit (2 g/litre) up to 16 g/litre. The brominated organic DBPs formed upon ozonation generally occur at low levels. The formation of chlorite can be estimated by a simple percentage (50-70%) of the applied chlorine dioxide dose. 6.2.2 Uncertainties of water quality data A toxicological study attempts to extrapolate a laboratory (controlled) animal response to a potential human response; one possible outcome is the estimation of cancer risk factors. An epidemiological study attempts to link human health effects (e.g., cancer) to a causative agent or agents (e.g., a DBP) and requires an exposure assessment. The chemical risks associated with disinfected drinking-water are potentially based on several routes of exposure: (i) ingestion of DBPs in drinking-water; (ii) ingestion of chemical disinfectants in drinking-water and the concomitant formation of DBPs in the stomach; and (iii) inhalation of volatile DBPs during showering. Although the in vivo formation of DBPs and the inhalation of volatile DBPs may be of potential health concern, the following discussion is based on the premise that the ingestion of DBPs present in drinking-water poses the most significant chemical health risk. Human exposure is a function of both DBP concentration and exposure time. More specifically, human health effects are a function of exposure to complex mixtures of DBPs (e.g., THMs versus HAAs, chlorinated versus brominated species) that can change seasonally/temporally (e.g., as a function of temperature and nature and concentration of NOM) and spatially (i.e., throughout a distribution system). Each individual chemical disinfectant can form a mixture of DBPs; combinations of chemical disinfectants can form even more complex mixtures. Upon their formation, most DBPs are stable, but some may undergo transformation by, for example, hydrolysis. In the absence of DBP data, surrogates such as chlorine dose (or chlorine demand), TOC (or UVA254) or bromide can be used to indirectly estimate exposure. While TOC serves as a good surrogate for organic DBP precursors, UVA254 provides additional insight into NOM characteristics, which can vary geographically. Two key water quality variables, pH and bromide, have been identified as significantly affecting the type and concentrations of DBPs that are produced. An exposure assessment should first attempt to define the individual types of DBPs and resultant mixtures likely to form as well as their time-dependent concentrations, as affected by their stability and transport through a distribution system. For epidemiological studies, some historical databases exist for disinfectant (e.g., chlorine) doses, possibly DBP precursor (e.g., TOC) concentrations, and possibly total THM (and in some cases, THM species) concentrations. In contrast to THMs, which have been monitored over longer time frames because of regulatory scrutiny, monitoring data for HAAs (and HAA species), bromate and chlorite are much more recent and hence sparse. However, DBP models can be used to simulate missing or past data (e.g., concentrations of HAAs can be predicted using data on THM concentrations). Another important consideration is documentation of past changes in water treatment practice. 6.2.3 Uncertainties of epidemiological data Even in well designed and well conducted analytical studies, relatively poor exposure assessments were conducted. In most studies, duration of exposure to disinfected drinking-water and the water source were considered. These exposures were estimated from residential histories and water utility or government records. In only a few studies was an attempt made to estimate a study participant's water consumption and exposure to either total THMs or individual THM species. In only one study was an attempt made to estimate exposures to other DBPs. In evaluating some potential risks, i.e., adverse outcomes of pregnancy, that may be associated with relatively short term exposures to volatile by-products, it may be important to consider the inhalation as well as the ingestion route of exposure from drinking-water. In some studies, an effort was made to estimate both by-product levels in drinking-water for etiologically relevant time periods and cumulative exposures. Appropriate models and sensitivity analysis such as Monte Carlo simulation can be used to help estimate these exposures for relevant periods. A major uncertainty surrounds the interpretation of the observed associations, as exposures to a relatively few water contaminants have been considered. With the current data, it is difficult to evaluate how unmeasured DBPs or other water contaminants may have affected the observed relative risk estimates. More studies have considered bladder cancer than any other cancer. The authors of the most recently reported results for bladder cancer risks caution against a simple interpretation of the observed associations. The epidemiological evidence for an increased relative risk of bladder cancer is not consistent -- different risks are reported for smokers and non-smokers, for men and women, and for high and low water consumption. Risks may differ among various geographic areas because the DBP mix may be different or other water contaminants are also present. More comprehensive water quality data must be collected or simulated to improve exposure assessments for epidemiological studies. 7. RISK CONCLUSIONS AND COMPARISONS Chlorination of drinking-water has been a cornerstone of efforts to prevent the spread of waterborne disease for almost a century (Craun et al., 1993). It is important to retain chlorination as an inexpensive and efficacious process unless a clear public health concern arises to eliminate it. It is uncertain that alternative chemical disinfectants reduce these estimated risks significantly (Bull & Kopfler, 1991). Identifying the safest way of producing drinking-water requires more conclusive toxicological or epidemiological evidence than is available today. It is important to recognize that there is a sizeable set of data already present on this issue and that resolution of this problem will not simply come from an expansion of that database. The focus must be elevated from questions of individual by-products and routine toxicological testing to a much more systematic approach towards the resolution of these larger issues. 7.1 Epidemiological studies The epidemiological associations between chlorinated drinking-water and human cancer have been subjected to several recent reviews, and the conclusions remain controversial. The small to medium relative risks for all the tumour sites studied (relative risks or odds ratios almost always less than 2) and uncertainty related to the magnitude and type of human exposures make it difficult to conclude that real risks result from the ingestion of chlorinated drinking-water. 7.2 Toxicological studies Toxicological studies are best suited for developing information on individual by-products or known combinations of by-products. The deficiencies in the present toxicological database are outlined below. 7.2.1 Diversity of by-products Significant qualitative and quantitative differences in the toxicological properties of DBPs have been demonstrated, depending upon whether they have some bromine substitution. Among the THMs, BDCM is of particular interest because it produces tumours in both rats and mice at several sites (NTP, 1987). Moreover, its potency calculated under the assumptions of the linearized multistage model is an order of magnitude greater than that of chloroform (Bull & Kopfler, 1991). DBCM produced liver tumours only in mice (NTP, 1985), but bromoform produced colon tumours in rats (NTP, 1989a). The fact that both BDCM and bromoform given in corn oil vehicle induce colon cancer in animals is of interest because of the epidemiological associations seen with colo-rectal tumours and consumption of chlorinated water. As with the THMs, however, a full complement of brominated and mixed bromochlorinated acetates are produced with the chlorination of drinking-water. These compounds have received little attention. While the chlorinated HAAs appear to be without significant genotoxic activity, the brominated HAAs appear to induce oxidative damage to DNA. Increases in the 8-OH-dG levels in hepatic DNA were observed with both acutely administered oral doses (Austin et al., 1996) and more prolonged exposures in drinking-water (Parrish et al., 1996). This activity increased with the degree of bromine substitution. Therefore, it cannot be concluded that the brominated HAAs are the mechanistic equivalents of the chlorinated HAAs. Association of mutagenic activity with the chlorination of drinking-water was first observed by Cheh et al. (1980). While some of the major DBPs are mutagenic, they are much too weak as mutagens to account for this activity. By far the largest individual contributor to this activity is the compound referred to as MX. This compound has been variously reported to account for up to 57% of the mutagenic activity produced in the chlorination of drinking-water (Meier et al., 1985a,b; Hemming et al., 1986; Kronberg & Vartiainen, 1988). MX has recently been shown to be a carcinogen in rats (Komulainen et al., 1997). As with other classes of DBPs, brominated analogues and structurally related compounds that could be of importance are produced in the chlorination of drinking-water (Daniel et al., 1991b; Suzuki & Nakanishi, 1995). 7.2.2 Diversity of modes of action It is important to recognize that the ways in which DBPs induce cancer are quite different. All of the modern work that has come forward on chloroform (Larson et al., 1994a,b, 1996) would strongly undermine the hypothesis that chloroform is contributing to the cancers observed in epidemiological studies. These toxicological results make a convincing case that tumorigenic responses in both the mouse liver and rat kidney are dependent upon necrosis and reparative hyperplasia. There is no basis for associating this type of target organ damage with the consumption of chlorinated drinking-water. On the other hand, brominated THMs are mutagenic (Zeiger, 1990; Pegram et al., 1997). It is generally assumed that mutagenic carcinogens will produce linear dose-response relationships at low doses, as mutagenesis is generally considered to be an irreversible and cumulative effect. In the HAA class, significant differences in mode of action have been demonstrated for DCA and TCA. Despite the close structural resemblance of DCA and TCA and their common target organ (liver cancer induction), it is becoming clear that the mechanisms by which they act are different. TCA is a peroxisome proliferator, and the tumour phenotype and genotype seen in mice are consistent with this being the mode of action by which it acts. However, DCA clearly produced tumours at doses below those that are required for peroxisome proliferation (DeAngelo et al., 1989, 1996; Daniel et al., 1992a; Richmond et al., 1995). The tumour phenotypes that DCA and TCA produce in mice are very different (Pereira, 1996; Pereira & Phelps, 1996; Stauber & Bull, 1997). From a risk assessment standpoint, however, one would question whether either DCA or TCA, alone, is likely to present significant cancer risk to humans at the low levels found in drinking-water. Neither compound appears to influence the carcinogenic process by a mutagenic mechanism (Stauber & Bull, 1997; Harrington-Brock et al., 1998; Stauber et al., 1998). Although different mechanisms appear to be involved, the mode of action for both compounds appears to be tumour promotion. In experiments of short duration, they tend to increase replication rates of normal hepatocytes; with more extended exposures or very high doses, however, they tend to depress replication rates (Carter et al., 1995; Stauber & Bull, 1997). There is evidence to suggest that the depression of cell replication is paralleled by depressed rates of apoptosis (Snyder et al., 1995). It is noteworthy that there is little support in the animal data for certain target organs that are prominently associated with chlorinated drinking-water in epidemiological studies (e.g., bladder cancer). Therefore, the possibility has to be left open that the carcinogenic effect of DBPs may be dependent on genetically determined characteristics of a target organ (or tissue) that make it more susceptible than the same organ in test animals. This problem can be resolved only by conducting toxicological studies in the appropriate human tissues and by developing much stronger epidemiological associations to guide these studies. The epidemiological studies can contribute to the resolution of the problem by (i) better identifying the drinking-water conditions that are associated with bladder or colo-rectal cancer, (ii) focusing on those characteristics of susceptibility that may increase the sensitivity at these target sites, and (iii) determining if interactions between biomarkers of susceptibility at these sites contribute to the epidemiological associations with disinfection of drinking-water. These "tasks" need to be accomplished sequentially. If epidemiological studies can provide insights into the first two tasks, then the experimental scientists can work much more profitably with epidemiologists to address the third task. 7.2.3 Reproductive, developmental and neurotoxic effects Much of this review has focused on questions related to chemical carcinogenesis, in part because that is where the bulk of the experimental data are found. There are other toxicological effects associated with some DBPs that could be of importance. Recently published epidemiological data (Waller et al., 1998) suggest the possibility that increased spontaneous abortion rates may be related to DBPs in drinking-water. Reproductive effects in females have been principally embryolethality and fetal resorptions associated with the HANs (Smith et al., 1988, 1989b). The dihaloacetates, DCA and DBA, have both been associated with effects on male reproduction, marked primarily by degeneration of the testicular epithelium (Toth et al., 1992; Linder et al., 1994a,b). Some effects on reproductive performance are noted at doses of DBA as low as 10 mg/kg of body weight per day. Dogs display testicular degeneration when administered doses of DCA of this same magnitude (Cicmanec et al., 1991). 7.3 Risks associated with mixtures of disinfectant by-products Disinfected drinking-water is a very complex mixture of chemicals, most of which have not been identified. Studies on individual DBPs may not represent the risk posed by the mixture. Research on complex mixtures was recently reviewed by ILSI (1998). Studies of simple combinations of chemicals provided positive results, but only at concentrations so much greater than those that occur in drinking-water as to be irrelevant. Studies utilizing complex mixtures of chemicals as they could be isolated from water or produced by chlorinating high concentrations of humic or fulvic acids produce little convincing evidence of adverse effect. A variety of methodological issues prevent our being too comfortable with that conclusion (ILSI, 1998). Moreover, the effort never developed to the point that the diverse qualities of water in various parts of the country could be taken into account. To be efficient, toxicological research needs to have a focus, and the research performed to solve this problem must be hypothesis-driven. This means that hypotheses of interactions would be based on knowledge of the toxic properties of individual by-products and would be subjected to experimental test. This can be much more efficient than designing complex multifactorial studies of all possible combinations of by-products produced by a disinfectant. An additional nuance would be to develop hypotheses as explicit tests of epidemiological findings. This would ensure that resources are appropriately directed and would provide a research agenda that would progress in a predictive way. At some stage, a hypothesis loses credibility or becomes recognized as being as close to "truth" as can be achieved experimentally. Similarly, epidemiological studies must begin to focus on what is known about the toxicology of individual DBPs. Testing of hypotheses about certain adverse health effects should begin with some understanding of which DBPs are known to produce an effect of interest in experimental systems. Epidemiologists then need to focus on those mechanisms of toxicity and interactions that are likely to be important at low doses (i.e., those that can be logically extrapolated to dose levels encountered from drinking disinfected drinking-water). Finally, information of this type should be used to develop new parameters that can be incorporated into the design of epidemiological studies. 8. CONCLUSIONS AND RECOMMENDATIONS Disinfection is unquestionably the most important step in the treatment of water for drinking-water supplies. The microbial quality of drinking-water should not be compromised because of concern over the potential long-term effects of disinfectants and DBPs. The risk of illness and death resulting from exposure to pathogens in drinking-water is very much greater than the risks from disinfectants and DBPs. Where local circumstances require that a choice be made between microbiological limits or limits for disinfectants and DBPs, the microbiological quality must always take precedence. Efficient disinfection must never be compromised. The microbiological quality of drinking-water is of paramount importance and must receive priority over any other considerations in relation to drinking-water treatment. However, the use of any chemical disinfectant results in the formation of by-products that themselves may be of health significance. A thorough understanding of how these DBPs form and the factors that control their formation is valuable in achieving a successful balance between satisfactory inactivation of pathogens and the minimization of DBP formation. The microbiological quality of drinking-water should always receive priority over the minimization of DBPs. Where it is possible, without compromising the microbiological quality of drinking-water, steps should be taken to minimize the concentrations of DBPs produced by the disinfectant(s) in use. Strategies to minimize exposure to DBPs should focus on the elimination of precursors through source water protection. Not only is this often the most efficient method of reducing DBP concentrations, but it will also assist in improving the microbiological quality of the water. Where treatment is required, DBP control strategies should emphasize DBP organic precursor (TOC) removal. 8.1 Chemistry Chlorine and alternative chemical disinfectants (ozone, chlorine dioxide and chloramine) all lead to the formation of DBPs. However, between disinfectants or combinations thereof, there are differences in DBP groups, species and mixtures that may affect human health. Key water quality determinants of DBPs include TOC, bromide and pH. Based on the current knowledge of both occurrence and health effects, the DBPs of most concern include total THMs and THM species, total HAAs and HAA species, bromate and chlorite. 8.2 Toxicology None of the chlorination by-products studied to date is a potent carcinogen at concentrations normally found in drinking-water. The toxicology of the DBPs suggests that the likelihood of adverse effects is not significantly different between the described disinfectant options. Toxicological information on mode and mechanism of action of disinfectants and their by-products is the major limitation for understanding the potential health risks at low doses. 8.3 Epidemiology Epidemiological studies have not identified an increased risk of cardiovascular disease associated with chlorinated or chloraminated drinking-water. The hypothesis of a causal relationship between consumption of chlorination by-products and the increased relative risk of any cancer remains an open question. There is insufficient epidemiological evidence to support a causal relationship between bladder cancer and exposures to chlorinated drinking-water, THMs, chloroform or other THM species. The epidemiological evidence is inconclusive and equivocal for an association between colon cancer and long duration of exposure to chlorinated drinking-water, THMs or chloroform. There is insufficient epidemiological information to properly interpret the observed risks for rectal cancer and the risks for other cancers observed in single analytical studies. The results of currently published studies do not provide convincing evidence that chlorinated water or THMs cause adverse pregnancy outcomes. 9. RESEARCH NEEDS 9.1 Chemistry of disinfectants and disinfectant by-products There is a need to consider the requirements of developing and developed countries. While future research needs are articulated below, technology transfer is important in implementing past research into practice. The disinfection practice common to all countries is chlorination; thus, chlorine DBPs should be the primary focus. * There is a need for a study to develop more DBP occurrence data from an international perspective; such an effort should also compile information on DBP precursors (TOC and bromide). * Because of the expertise required to measure certain DBPs (e.g., HAAs) and the deficiencies of historical databases, there is a need to develop improved models for predicting DBP formation and precursor removal, allowing the use of DBPs (e.g., chloroform) or surrogates (e.g., TOC) that are more simple to measure. These models can also be used to predict the formation of DBPs from the use of a particular disinfectant and the factors that control the appearance and formation of these DBPs, thus allowing appropriate control strategies to be developed and better assessment of exposure. * Although HAAs have been monitored for several years, this monitoring effort has been based on measurements of the sum of five or six species. There is a need to develop more information on the occurrence of the nine species of HAAs. * For analytical reasons, non-polar organic by-products (measurable by GC) and ionic by-products (measurable by IC) have received more attention. There is a need to develop analytical methods for polar by-products and to define their occurrence. * A significant percentage of TOX remains unaccounted for by specific halogenated DBPs; there is a need to identify these compounds. * Given the health effects data on bromate and the anticipated higher ozone doses that will be required to inactivate Cryptosporidium, more data are needed on bromate formation in low-bromide waters when high ozone doses are used. * More information is needed on the composition of NOM to assist in determining the type and extent of DBP formation that can be expected in a given water and how well treatment processes will achieve precursor removal. * There is a need to develop a better understanding of how water quality parameters affect the extent of bromine incorporation into DBPs. * To assist small water supply systems in minimizing DBP formation, there is a need to develop simple, easily operated treatment systems for the removal of NOM from source waters. 9.2 Toxicology * Toxicological research needs to be focused to be effective. One approach to accomplish this is to design experiments to determine if a biological basis can be established for the epidemiological findings. * Further toxicological characterization of the effects of DBPs at low doses is needed. These studies should be directed at understanding the mechanisms that are operative at these doses in humans. * There are at least three areas that need the attention of both toxicologists and epidemiologists: risks posed by brominated by-products, how risks are modified by pH, and risks posed by the use of a disinfectant as an oxidizing agent during drinking-water treatment. * A relatively small fraction of DBPs has received substantive toxicological study. Future studies should be directed at those by-products that occur with high frequency and at relatively high concentrations. * Little attention has been paid to those individuals in a population who possess sensitivities to particular chemicals and/or modes of action because of genetic and lifestyle factors. An example is major polymorphic differences in enzymes that metabolize DBPs. * Humans are exposed to complex mixtures of disinfectants and DBPs. It is becoming apparent that chemicals with like mechanisms interact in an additive way at low concentrations. Little information exists on the potential non-additive interaction of chemicals with different mechanisms. 9.3 Epidemiology * Additional studies to evaluate cancer risks should be analytical and should include a more comprehensive assessment of drinking-water exposures, especially for DBPs, for etiologically relevant time periods. The interpretation of the results from currently conducted studies of both cancer and adverse pregnancy outcomes suffers from lack of knowledge about exposures to other DBPs and other water contaminants. It may be possible to assess additional water exposures for study participants in several recently conducted studies of cancer and reproductive risks. If this can be done, a better estimate of exposure to water contaminants and DBPs will be available for additional analyses of risks in these study populations. Depending on the results of these reanalyses, the need for additional studies of possible cancer risks can be better evaluated. * It is important to improve the quality of future epidemiological studies with adequate and appropriate exposure information. This underscores the need to include in the planning and conduct of epidemiological studies individuals who are knowledgeable about DBP chemistry. * Additional studies are needed to better assess the importance of the observed association between DBPs and early miscarriage or neural tube defects. Research should continue on possible reproductive and developmental effects associated with drinking-water disinfection. * The most important research need is to improve the assessment of drinking-water exposures for epidemiological studies. There is a need for improved models that can be used to estimate exposures to various specific by-products and mixtures of by-products. There is also a need to collect more complete information about individual water consumption and activity patterns that may influence exposure assessments. Better and more complete exposure information will improve the sensitivity of epidemiological studies. 10. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES A number of disinfectants and DBPs have been considered by IARC. On the basis of the available published data, the most recent classification of these chemicals is as follows: Disinfectants Hypochlorite salts: Group 3 (1991) Disinfectant by-products Trihalomethanes Bromodichloromethane: Group 2B (1999) Dibromochloromethane: Group 3 (1999) Bromoform: Group 3 (1999) Haloacetic acids Dichloroacetic acid: Group 3 (1995) Trichloroacetic acid: Group 3 (1995) Haloacetaldehyde Chloral and chloral hydrate: Group 3 (1995) Haloacetonitriles Bromochloroacetonitrile: Group 3 (1999) Chloroacetonitrile: Group 3 (1999) Dibromoacetonitrile: Group 3 (1999) Dichloroacetonitrile: Group 3 (1999) Trichloroacetonitrile: Group 3 (1999) Other disinfectant by-products Potassium bromate: Group 2B (1987) Sodium chlorite: Group 3 (1991) In addition, IARC has classified chlorinated drinking-water in Group 3 (1991). Disinfectants and DBPs were evaluated in the Guidelines for drinking-water quality, and the following guideline values recommended (WHO, 1993, 1996, 1998): Disinfectants Chlorine (hypochlorous acid and hypochlorite): 5 mg/litre (1993) Monochloramine: 3 mg/litre (1993) Disinfectant by-products Trihalomethanes Bromodichloromethane: 60 g/litre for an excess lifetime cancer risk of 10-5 (1993) Dibromochloromethane: 100 g/litre (1993) Bromoform: 100 g/litre (1993) Chloroform: 200 g/litre (1998) Haloacetic acids Dichloroacetic acid: 50 g/litre (provisional) (1993) Trichloroacetic acid: 100 g/litre (provisional) (1993) Haloacetaldehyde Chloral hydrate: 10 g/litre (provisional) (1993) Haloacetonitriles Dibromoacetonitrile: 100 g/litre (provisional) (1993) Dichloroacetonitrile: 90 g/litre (provisional) (1993) Trichloroacetonitrile: 1 g/litre (provisional) (1993) Other disinfectant by-products Bromate: 25 g/litre (provisional) for an excess lifetime cancer risk of 7 10-5 (1993) Chlorite: 200 g/litre (provisional) (1993) Cyanogen chloride (as cyanide): 70 g/litre (1993) Formaldehyde: 900 g/litre (1993) 2,4,6-Trichlorophenol: 200 g/litre for an excess lifetime cancer risk of 10-5 (1993) JECFA (FAO/WHO, 1993) evaluated potassium bromate and concluded that it was genotoxic and carcinogenic. The International Programme on Chemical Safety (IPCS) published an evaluation of chloroform in its Environmental Health Criteria Monograph series (WHO, 1994). The IPCS Concise International Chemical Assessment Document (CICAD) on chloral hydrate (in press) considers the dose level of 16 mg/kg in the study of Sanderson et al. (1982) to be a NOAEL rather than a LOAEL (as in the present document), and therefore uses studies in humans as the basis of a tolerable intake. On the basis of a LOAEL of 11 mg/kg, and using an uncertainty factor of 10 for intraspecies variation and 10 for conversion of LOAEL to NOAEL, the CICAD derived a tolerable intake of 0.1 mg/kg. REFERENCES Abbas R & Fisher JW (1997) A physiologically-based pharmacokinetic model for trichloroethylene and its metabolites, chloral hydrate, trichloroacetate, dichloroacetate, trichloroethanol and trichloroethanol glucuronide in B6C3F1 mice. Toxicol Appl Pharmacol, 147: 15-30. Abbas RR, Seckel CS, Kidney JK, & Fisher JW (1996) Pharmacokinetic analysis of chloral hydrate and its metabolism in B6C3F1 mice. Drug Metab Dispos, 24: 1340-1346. Abdel-Rahman MS, Couri D, & Bull RJ (1980) Kinetics of ClO2 and effects of ClO2, ClO2-, and ClO3- in drinking water on blood glutathione and hemolysis in rat and chicken. J Environ Pathol Toxicol, 3: 431-449. Abdel-Rahman MS, Berardi MR, & Bull RJ (1982a) Effect of chlorine and monochloramine in drinking water on the developing rat fetus. J Appl Toxicol, 2(3): 156. Abdel-Rahman MS, Couri D, & Bull RJ (1982b) Metabolism and pharmacokinetics of alternate drinking water disinfectants. Environ Health Perspect, 46: 19-23. Abdel-Rahman MS, Waldron DM, & Bull RJ (1983) A comparative kinetics study of monochloramine and hypochlorous acid in rat. J Appl Toxicol, 3: 175-179. Abdel-Rahman MS, Couri D, & Bull RJ (1984a) Toxicity of chlorine dioxide in drinking water. J Am Coll Toxicol, 3: 277-284. Abdel-Rahman MS, Couri D, & Bull RJ (1984b) The kinetics of chlorite and chlorate in the rat. J Am Coll Toxicol, 3: 261-267. Acharya S, Mehta K, Rodrigues S, Pereira J, Krishnan S, & Rao CV (1995) Administration of subtoxic doses of t-butyl alcohol and trichloroacetic acid to male Wistar rats to study the interactive toxicity. Toxicol Lett, 80: 97-104. Ade P, Guastadisegni C, Testai E, & Vittozzi L (1994) Multiple activation of chloroform in kidney microsomes from male and female DBA/2J mice. J Biochem Toxicol, 9: 289-295. Adler ID (1993) Synopsis of the in vivo results obtained with the 10 known or suspected aneugens tested in the CEC collaborative study. Mutat Res, 287(1): 131-137. Agarwal AK & Mehendele HM (1983) Absence of potentiation of bromoform hepatotoxicity and lethality by chlordecone. Toxicol Lett, 15: 251-257. Ahmed AE, Soliman SA, Loh JP, & Hussein GI (1989) Studies on the mechanism of haloacetonitriles toxicity: Inhibition of rat hepatic glutathione- S-transferases in vitro. Toxicol Appl Pharmacol, 100: 271-279. Ahmed AE, Jacob S, & Loh JP (1991) Studies on the mechanism of haloacetonitriles toxicity: quantitative whole body autoradiographic distribution of [2-14C]chloroacetonitrile in rats. Toxicology, 67: 279-302. Aida Y, Takada K, Uchida O, Yasuhara K, Kurokawa Y, & Tobe M (1992a) Toxicities of microencapsulated tribromomethane, dibromochloromethane and bromodichloromethane administered in the diet to Wistar rats for one month. J Toxicol Sci, 17: 119-133. Aida Y, Yasuhara K, Takada K, Kurokawa Y, & Tobe M (1992b) Chronic toxicity of microencapsulated bromodichloromethane administered in the diet to Wistar rats. J Toxicol Sci, 17: 51-68, [Erratum] 17:167. Aieta EM & Berg JD (1986) A review of chlorine dioxide in drinking water treatment. J Am Water Works Assoc, 78(6): 62-72. Alavanja M, Goldstein I, & Susser M (1978) Case-control study of gastrointestinal and urinary tract cancer mortality and drinking water chlorination. In: Jolley RL, Gorchev H, & Hamilton DH ed. Water chlorination: Environmental impact and health effects. Ann Arbor, Michigan, Ann Arbor Science Publishers, vol 2, pp 395-409. Allen JW, Collins BW, & Evansky PA (1994) Spermatid micronucleus analyses of trichloroethylene and chloral hydrate effects in mice. Mutat Res, 323(1-2): 81-88. Almeyda J & Levantine A (1972) Cutaneous reactions to barbiturates, chloral hydrate and its derivatives. Br J Dermatol, 86: 313-316. Ames RG & Stratton JW (1987) Effect of chlorine dioxide water disinfection on hematologic and serum parameters of renal dialysis. Arch Environ Health, 42(5): 280-285. Ammann P, Laethem CL, & Kedderis GL (1998) Chloroform-induced cytolethality in freshly isolated male B6C3F1 mouse and F344 rat hepatocytes. Toxicol Appl Pharmacol, 149(2): 217-225. Amy G, Chadik PA, & Chowdhury ZK (1987) Developing models for predicting THM formation potential and kinetics. J Am Water Works Assoc, 79(7): 89-96. Amy G, Siddiqui M, Ozekin K, & Westerhoff P (1993) Threshold levels for bromate ion formation in drinking water. In: Proceedings of the International Water Supply Association Workshop, Paris. Boston, Massachusetts, Blackwell Scientific Publications, pp 169-180. Amy G, Siddiqui M, Zhai W, & Debroux J (1994) Survey on bromide in drinking water and impacts on DBP formation. Denver, Colorado, American Water Works Association (Report No. 90662). Amy G, Siddiqui M, Ozekin K, Zhu HW, & Wang C (1998) Empirically-based models for predicting chlorination and ozonation by-products: Trihalomethanes, haloacetic acids, chloral hydrate, and bromate. Cincinnati, Ohio, US Environmental Protection Agency (EPA-815-R-98-005). Anders MW, Stevens JL, Sprague RW, Shaath Z, & Ahmed AE (1978) Metabolism of haloforms to carbon monoxide. II. In vivo studies. Drug Metab Dispos, 6: 556-560. Andrews SA & Ferguson MJ (1995) Minimizing DBP formation while ensuring Giardia control. In: Minear RA & Amy GL ed. Disinfection by-products in water treatment. Chelsea, Michigan, Lewis Publishers, Inc. Andrews SA, Huch PM, Chute AJ, Bolton JR, & Anderson WA (1996) UV oxidation for drinking water-feasibility studies for addressing specific water quality issues. In: Proceedings of the Water Quality Technology Conference, New Orleans, LA. Denver, Colorado, American Water Works Association. Angel P & Karin M (1991) The role of Jun, Fos, and the AP-1 complex in cell proliferation and transformation. Biochim Biophys Acta, 1072: 126-157. Anna CH, Maronpot RR, Pereira MA, Foley JF, Malarkey DE, & Anderson MW (1994) Ras proto-oncogene activation in dichloroacetic-, trichloroethylene- and tetrachloroethylene-induced liver tumors in B6C3F1 mice. Carcinogenesis, 15: 2255-2261. APHA (American Public Health Association) (1995) Standard methods for the examination of water and wastewater, 19th ed. Washington, DC, American Public Health Association/American Water Works Association/Water Pollution Control Federation. Aschengrau A, Zierler S, & Cohen A (1989) Quality of community drinking water and the occurrence of spontaneous abortions. Arch Environ Health, 44(5): 283-290. Aschengrau A, Zierler S, & Cohen A (1993) Quality of community drinking water and the occurrence of late adverse pregnancy outcomes. Arch Environ Health, 48(2): 105-114. Ashby J, Mohammed R, & Callander RD (1987) N-Chloropiperidine and calcium hypochlorite: Possible examples of toxicity-dependent clastogenicity, in vitro. Mutat Res, 189: 59-68. Austin EW & Bull RJ (1997) The effect of pretreatment with dichloroacetate and trichloroacetate on the metabolism of bromodichloroacetate. J Toxicol Environ Health, 52: 367-383. Austin EW, Okita JR, Okita RT, Larson JL, & Bull RJ (1995) Modification of lipoperoxidative effects of dichloroacetate and trichloroacetate is associated with peroxisome proliferation. Toxicology, 97: 59-69. Austin EW, Parrish JM, Kinder DH, & Bull RJ (1996) Lipid peroxidation and formation of 8-hydroxydeoxyguanosine from acute doses of halogenated acetic acids. Fundam Appl Toxicol, 31: 77-82. AWWARF (1991) Disinfection by-products database and model project. Denver, Colorado, American Water Works Association Research Foundation. Bailar JC (1995) The practice of meta-analysis. J Clin Epidemiol, 48(1): 149-157. Bailey PS (1978) Ozonation in organic chemistry. New York, London, Academic Press. Balster RL & Borzelleca JF (1982) Behavioral toxicity of trihalomethane contaminants of drinking water in mice. Environ Health Perspect, 46: 127-136. Banerji AP & Fernandes AO (1996) Field bean protease inhibitor mitigates the sister-chromatid exchanges induced by bromoform and depresses the spontaneous sister-chromatid exchange frequency of human lymphocytes in vitro. Mutat Res, 360: 29-35. Bean JA, Isacson P, Hausler WJ, & Kohler J (1982) Drinking water and cancer incidence in Iowa I. Trends and incidence by source of drinking water and size of municipality. Am J Epidemiol, 116(6): 912-923. Bempong MA & Scully FE Jr (1985) Mutagenicity and clastogenicity of N-chloropiperidine. J Environ Pathol Toxicol, 6: 241-251. Bempong MA, Montgomery C, & Scully FE Jr (1980) Mutagenic activity of N-chloropiperidine. J Environ Pathol Toxicol, 4: 345-354. Bercz JP & Bawa R (1986) Iodination of nutrients in the presence of chlorine based disinfectants used in drinking water treatment. Toxicol Lett, 34(2-3): 141-147. Bercz JP, Jones L, Garner L, Murray D, Ludwig A, & Boston J (1982) Subchronic toxicity of chlorine dioxide and related compounds in drinking water in the nonhuman primate. Environ Health Perspect, 46: 47-55. Bercz JP, Jones LL, Harrington RM, Bawa R, & Condie L (1986) Mechanistic aspects of ingested chlorine dioxide on thyroid function: Impact of oxidants on iodide metabolism. Environ Health Perspect, 69: 249-255. Bersin RM, Wolfe C, Kwasman M, Lau D, Klinski C, Tanaka K, Khorrami, P, Henderson GN, De Marco T, & Chatterjee K (1994) Improved hemodynamic function and mechanical efficiency in congestive heart failure with sodium dichloroacetate. J Am Coll Cardiol, 23: 1617-1624. Bhat HK, Kanz MF, Campbell GA, & Ansari GAS (1991) Ninety day toxicity study of chloroacetic acids in rats. Fundam Appl Toxicol, 17: 240-253. Bhunya SP & Behera BC (1987) Relative genotoxicity of trichloroacetic acid (TCA) as revealed by different cytogenetic assays: Bone marrow chromosome aberration, micronucleus and sperm-head abnormality in the mouse. Mutat Res, 188: 215-221. Bignami M, Conti G, Conti L, Crebelli R, Misuraca F, Puglia AM, Randazzo R, Sciandrello G, & Carere A (1980) Mutagenicity of halogenated aliphatic hydrocarbons in Salmonella typhimurium, Streptomyces coelicolor and Aspergillus nidulans. Chem-Biol Interact, 30: 9-23. Blackshear PJ, Holloway PA, & Alberti KM (1974) The metabolic effects of sodium dichloroacetate in the starved rat. Biochem J, 142: 279-286. Blazak WF, Meier JR, Stewart BE, Blachman DC, & Deahl JT (1988) Activity of 1,1,1- and 1,1,3-trichloroacetones in a chromosomal aberration assay in CHO cells and the micronucleus and spermhead abnormality assays in mice. Mutat Res, 206: 431-438. Bloxham CA, Wright N, & Hoult JG (1979) Self-poisoning by sodium chlorate: Some unusual features. Clin Toxicol, 15: 185-188. Bolyard M & Fair PS (1992) Occurrence of chlorate in hypochlorite solutions used for drinking water treatment. Environ Sci Technol, 26(8): 1663-1665. Bolyard M, Fair PS, & Hautman DP (1993) Sources of chlorate ion in U.S. drinking water. J Am Water Works Assoc, 85: 81-88. Borzelleca JF & Carchman RA (1982) Effects of selected organic drinking water contaminants on male reproduction. Research Triangle Park, North Carolina, US Environmental Protection Agency (EPA 600/1-82-009; NTIS PB82-259847). Bove FJ, Fulcomer MC, Koltz JB, Esmart J, Dufficy EM, & Zagraniski RT (1992a) Report on phase IV-A: public drinking water contamination and birthweight fetal deaths, and birth defects, a cross-sectional study. Trenton, New Jersey, New Jersey Department of Health. Bove FJ, Fulcomer MC, Koltz JB, Esmart J, Dufficy EM, Zagraniski RT, & Savrin JE (1992b) Report on phase IV-B: public drinking water contamination and birthweight and selected birth defects, a case-control study. Trenton, New Jersey, New Jersey Department of Health. Bove FJ, Fulcomer MC, Klotz JB, Esmart J, Dufficy EM, & Savrin JE (1995) Public drinking water contamination and birth outcomes. Am J Epidemiol, 141(9): 850-862. Bowman FJ, Borzelleca JF, & Munson AE (1978) The toxicity of some halomethanes in mice. Toxicol Appl Pharmacol, 44: 213-215. Brennan RJ & Schiestl RH (1998) Chloroform and carbon tetrachloride induce intra-chromosomal recombination and oxidative free radicals in Saccharomyces cerevisiae. Mutat Res, 397(2): 271-278. Brenniman GR, Vasilomanolakis-Lagos J, Amsel J, Namekata T, & Wolff AH (1980) Case-control study of cancer deaths in Illinois communities served by chlorinated or nonchlorinated water. In: Jolley RL, Brungs WA, Cumming RB, & Jacobs VA ed. Water chlorination: Environmental impact and health effects. Ann Arbor, Michigan, Ann Arbor Science Publishers, Inc., vol 3, pp 1/043-1/057. Bronzetti G, Galli A, Corsi C, Cundari E, Del Carratore R, Nieri R, & Paolini M (1984) Genetic and biochemical investigation on chloral hydrate in vitro and in vivo. Mutat Res, 141: 19-22. Brown DM, Langley PF, Smith D, & Taylor DC (1974) Metabolism of chloroform - 1. The metabolism of [14C]chloroform by different species. Xenobiotica, 4: 151-163. Brown-Woodman PD, Hayes LC, Huq F, Herlihy C, Picker K, & Webster WS (1998) In vitro assessment of the effect of halogenated hydrocarbons: Chloroform, dichloromethane, and dibromoethane on embryonic development of the rat. Teratology, 57(6): 321-333. Bruchet A, Costentin E, Legrand MF, & Malleviale J (1992) Influence of the chlorination of natural nitrogenous organic compounds on tastes and odours in finished drinking waters. Water Sci Technol, 25(2): 323-333. Brunborg G, Holme JA, Soderlund EJ, & Dybing E (1990) Organ-specific genotoxic effects of chemicals: The use of alkaline elution to detect DNA damage in various organs of in vivo exposed animals. Prog Clin Biol RES, 340D: 43-52. Brunborg G, Holme JA, Soderlund EJ, Hongslo JK, Vartiainen T, Lotjonen S, & Becher G (1991) Genotoxic effects of the drinking water mutagen 3-chloro-4-(dichloromethyl)-5-hydroxy-2[5H]-furanone (MX) in mammalian cells in vitro and in rats in vivo. Mutat Res, 260: 55-64. Bruschi SA & Bull RJ (1993) In vitro cytotoxicity of mono-, di- and trichloroacetate and its modulation by peroxisome proliferation. Fundam Appl Toxicol, 21: 366-375. Brusick D (1986) Genotoxic effects in cultured mammalian cells produced by low pH treatment conditions and increased ion concentrations. Environ Mutagen, 8: 879-886. Bull RJ (1980) Health effects of alternate disinfectants and their reaction products. J Am Water Works Assoc, 72: 299-303. Bull RJ (1982a) Health effects of drinking water disinfectants and disinfectant by-products. Environ Sci Technol, 16: 554A-559A. Bull RJ (1982b) Toxicological problems associated with alternative methods of disinfection. J Am Water Works Assoc, 74: 642-648. Bull RJ (1992) Toxicology of drinking water disinfection. In: Lippman M ed. Environmental toxicants: Human exposures and their health effects. New York, Van Nostrand Reinhold, pp 184-230. Bull RJ (1993) Toxicology of disinfectants and disinfectant by-products. In: Craun GF ed. Safety of water disinfection: Balancing chemical and microbial risks. Washington, DC, ILSI Press, pp 239-256. Bull RJ & DeAngelo AB (1995) Carcinogenic properties of brominated haloacetates - Disinfection by-products in drinking water: Critical issues in health effects research. Washington, DC, International Life Sciences Institute, p 29. Bull RJ & Robinson M (1985) Carcinogenic activity of haloacetonitrile and haloacetone derivatives in the mouse skin and lung. In: Jolley RL, Bull RJ, & Davis WP ed. Water chlorination: Chemistry, environmental impact and health effects. Chelsea, Michigan, Lewis Publishers, Inc., vol 5, pp 221-236. Bull RJ & Kopfler FC (1991) Health effects of disinfectants and disinfection by-products. Denver, Colorado, American Water Works Association Research Foundation and American Water Works Association. Bull RJ, Meier JR, Robinson M, Ringhand HP, Laurie RD, & Stober JA (1985) Evaluation of mutagenic and carcinogenic properties of brominated and chlorinated haloacetonitriles: By-products of chlorination. Fundam Appl Toxicol, 5: 1065-1074. Bull RJ, Sanchez IM, Nelson MA, Larson JL, & Lansing AL (1990) Liver tumor induction in B6C3F1 mice by dichloroacetate and trichloroacetate. Toxicology, 63: 341-359. Bull RJ, Templin M, Larson JL, & Stevens DK (1993) The role of dichloroacetate in the hepatocarcinogenicity of trichloroethylene. Toxicol Lett, 68: 203-211. Bull RJ, Birnbaum LS, Cantor KP, Rose JB, Butterworth BE, Pegram R, & Tuomisto J (1995) Water chlorination: Essential process or cancer hazard? Fundam Appl Toxicol, 28: 155-166. Burton-Fanning FW (1901) Poisoning by bromoform. Br Med J, 18 May: 1202-1203. Butler TC (1948) The metabolic fate of chloral hydrate. J Pharmacol Exp Ther, 92: 49-58. Butler TC (1949) Reduction and oxidation of chloral hydrate by isolated tissues in vitro. J Pharmacol Exp Ther, 95: 360-362. Butterworth BE, Templin MV, Constan AA, Sprankle CS, Wong BA, Pluta LJ, Everitt JI, & Recio L (1998) Long-term mutagenicity studies with chloroform and dimethylnitrosamine in female lacI transgenic B6C3F1 mice. Environ Mol Mutagen, 31(3): 248-256. Cabana BE & Gessner PK (1970) The kinetics of chloral hydrate metabolism in mice and the effect thereon of ethanol. J Pharmaocol Exp Ther, 174: 260-275. Canada (1993) Guidelines for Canadian drinking water quality. Ottawa, Ontario, Health Canada, Health Protection Branch. Cantor KP (1997) Drinking water and cancer. Cancer Causes Control, 8(3): 292-308. Cantor KP, Hoover R, & Hartge P (1985) Drinking water source and bladder cancer: A case-control study. In: Jolley RL, Bull RJ, & Davis WP ed. Water chlorination: Chemistry, environmental impact, and health effects. Chelsea, Michigan, Lewis Publishers, Inc., vol 5, pp 145-152. Cantor KP, Hoover R, & Hartge P (1987) Bladder cancer, drinking water source, and tap water consumption: A case-control study. J Natl Cancer Inst, 79(1): 1269-1279. Cantor KP, Hoover R, & Hartge P (1990) Bladder cancer, tap water consumption, and drinking water source. In: Jolley RL, Condie LW, & Johnson JD ed. Water chlorination: Chemistry, environmental impact, and health effects. Chelsea, Michigan, Lewis Publishers, Inc., vol 6, pp 411-419. Cantor KP, Lynch CF, & Hildesheim M (1995) Chlorinated drinking water and risk of bladder, colon, and rectal cancers: A case-control study in Iowa. In: Disinfection by-products in drinking water: Critical issues in health effects research. Washington, DC, International Life Sciences Institute, p 133. Cantor KP, Lynch CF, & Hildesheim M (1996) Chlorinated drinking water and risk of glioma: a case-control study in Iowa, USA. Epidemiology, 7(suppl 4): S83. Cantor KP, Lynch CF, Hildesheim ME, Dosemeci M, Lubin J, Alavanja M, & Craun G (1998) Drinking water source and chlorination byproducts: Risk of bladder cancer. Epidemiology, 9(1): 21-28. Carlo GL & Mettlin CJ (1980) Cancer incidence and trihalomethane concentrations in a public drinking water system. Am J Public Health, 70: 523-525. Carlson M & Hardy D (1998) Controlling DBPs with monochloramine. Effect of water quality conditions on controlling disinfection by-products with chloramines. J Am Water Works Assoc, 90(2): 95-106. Carlton BD, Bartlett A, Basaran AH, Colling K, Osis I, & Smith MK (1986) Reproductive effects of alternative disinfectants. Environ Health Perspect, 69: 237-241. Carlton BD, Habash DL, Basaran AH, George EL, & Smith MK (1987) Sodium chlorite administration in Long-Evans rats: Reproductive and endocrine effects. Environ Res, 42: 238-245. Carlton BD, Basaran AH, Mezza LE, George EL, & Smith MK (1991) Reproductive effects in Long-Evans rats exposed to chlorine dioxide. Environ Res, 56: 170-177. Carraro F, Klein S, Rosenblatt JI, & Wolfe RR (1989) The effect of dichloroacetate on lactate concentration in exercising humans. J Appl Physiol, 66: 591-597. Carter JH, Carter HW, & DeAngelo AB (1995) Biochemical, pathologic and morphometric alterations induced in male B6C3F1 mouse liver by short-term exposure to dichloroacetic acid. Toxicol Lett, 81: 55-71. Cattley RC, DeLuca J, Elcombe C, Fenner-Crisp P, Lake BG, Marsman DS, Pastoor TA, Popp JA, Robinson DE, Schwetz B, Tugwood J, & Wahli W (1998) Do peroxisome proliferating compounds pose a hepatocarcinogenic hazard to humans? Regul Toxicol Pharmacol, 27: 47-60. Cavanagh JE, Weinberg HS, Gold A, Sangalah R, Marbury D, & Glaze WH (1992) Ozonation by-products: identification of bromohydrins from the ozonation of natural waters with enhanced bromide levels. Environ Sci Technol, 26(8): 1658-1662. Chang J-HS, Vogt CR, Sun GY, & Sun AY (1981) Effects of acute administration of chlorinated water on liver lipids. Lipids, 16: 336-340. Chang LW, Daniel FB, & DeAngelo AB (1992) Analysis of DNA strand breaks induced in rodent liver in vivo, hepatocytes in primary culture, and a human cell line by chlorinated acetic acids and chlorinated acetaldehydes. Environ Mol Mutagen, 20: 277-288. Cheh AM, Skochdopole J, Koski P, & Cole L (1980) Nonvolatile mutagens in drinking water: Production by chlorination and destruction by sulfite. Science, 207: 90-92. Chipman JK, Davies JE, Parson JL, Mair J, O'Neil G, & Fawell JK (1998) DNA oxidation by potassium bromate: a direct mechanism or linked to lipid peroxidation? Toxicology, 126(2): 93-102. Cho DH, Hong JT, Chin K, Cho TS, & Lee BM (1993) Organotropic formation and disappearance of 8-hydroxydeoxyguanosine in the kidney of Sprague-Dawley rats exposed to Adriamycin and KBrO3. Cancer Lett, 74: 141-145. Chow BM & Roberts PV (1981) Halogenated by-product formation by chlorine dioxide and chlorine. J Environ Eng Div, 107: 609-618. Christ SA, Read EJ, Stober JA, & Smith MK (1996) Developmental effects of trichloroacetonitrile administered in corn oil to pregnant Long-Evans rats. J Toxicol Environ Health, 47: 233-247. Christman RF, Norwood DS, Millington DS, & Johnson JD (1983) Identity and yields of major halogenated products of aquatic fulvic acid chlorination. Environ Sci Technol, 17: 625-628. Chu I, Secours VE, Marino I, & Villeneuve DC (1980) The acute toxicity of four trihalomethanes in male and female rats. Toxicol Appl Pharmacol, 5: 351-353. Chu I, Villeneuve DC, Secours VE, Becking GC, & Valli VE (1982) Toxicity of trihalomethanes: I. The acute and subacute toxicity of chloroform, bromodichloromethane, chlorodibromomethane, and bromoform in rats. J Environ Sci Health, B17: 205-224. Cicmanec JL, Condie LW, Olson GR, & Wang SR (1991) 90-day toxicity study of dichloracetate in dogs. Fundam Appl Toxicol, 17: 376-389. Clark RM, Adams JQ, & Kyjins BW (1994) DBP control in drinking water: cost and performance. J Environ Eng, 120(4): 759-771. Claude J, Kunze E, & Frentzel-Beyne R (1986) Life-style and occupational risk factors in cancer of the lower urinary tract. Am J Epidemiol, 124: 578. Claus TH & Pilkis SJ (1977) Effect of dichloroacetate and glucagon on the incorporation of labeled substrates into glucose and on pyruvate dehydrogenase in hepatocytes from fed and starved rats. Arch Biochem Biophys, 182: 52-63. CMA (1997) Sodium chlorite: drinking water rat two-generation reproductive toxicity study. Washington, DC, Chemical Manufacturers Association (Quintiles Report CMA/17/96). Cole WJ, Mitchell RG, & Salamonsen RF (1975) Isolation, characterization and quantitation of chloral hydrate as a transient metabolite of trichloroethylene in man using electron capture gas chromatography and mass fragmentography. J Pharm Pharmacol, 27: 167-171. Coleman WE, Munch JW, Kaylor WH, Streicher RP, Ringhand HP, & Meier JR (1984) Gas chromatography/mass spectroscopy analysis of mutagenic extracts of aqueous chlorinated humic acid. A comparison of the byproducts to drinking water contaminants. Environ Sci Technol, 18: 674-681. Coleman WE, Munch JW, & Kopfler FC (1992) Ozonation/post-chlorination of humic acids: a model for predicting drinking water DBPs. J Ozone Sci Eng, 14: 349-355. Condie LW, Smallwood CL, & Laurie RD (1983) Comparative renal and hepatotoxicity of halomethanes: bromodichloromethane, bromoform, chloroform, dibromochloromethane and methylene chloride. Drug Chem Toxicol, 6: 563-578. Connor MJ & Gillings D (1974) An empiric study of ecological inference. Am J Public Health, 74: 555-559. Connor PM, Moore GS, Calabrese EJ, & Howe GR (1985) The renal effects of sodium chlorite in the drinking water of C57L/J male mice. J Environ Pathol Toxicol Oncol, 6: 253-260. Conolly RB & Butterworth BE (1995) Biologically based dose response model for hepatic toxicity: a mechanistically based replacement for traditional estimates of noncancer risk. Toxicol Lett, 82: 901-906. Cooper WJ, Meyer LM, Bofill CC, & Cordal E (1983) Quantitative effects of bromine on the formation and distribution of trihalomethanes in groundwater with a high organic content. In: Jolley RL, Brungs WA, Cotruvo JA, Cumming RB, Mattice JS, & Jacobs VA ed. Water chlorination: Environmental impacts and health effects - Book 1: Chemistry and water treatment. Ann Arbor, Michigan, Ann Arbor Science Publishers, vol 4, pp 285-296. Cooper WJ, Zika RG, & Steinhauer MS (1985) Bromide-oxidant interactions and THM formation: a literature review. J Am Water Works Assoc, 77(4): 116-121. Corley RA, Mendrala AL, Smith FA, Staats DA, Gargas ML, Conolly RB, Andersen ME, & Reitz RH (1990) Development of a physiologically based pharmacokinetic model for chloroform. Toxicol Appl Pharmacol, 103: 512-527. Cosby NC & Dukelow WR (1992) Toxicology of maternally ingested trichloroethylene (TCE) on embryonal and fetal development in mice and of TCE metabolites on in vitro fertilization. Fundam Appl Toxicol, 19: 268-274. Cotter JL, Fader RC, Lilley C, & Herndon DN (1985) Chemical parameters, antimicrobial activities, and tissue toxicity of 0.1 and 0.5% sodium hypochlorite solutions. Antimicrob Agents Chemother, 28: 118-122. Coude FX, Saudubray JM, DeMaugre F, Marsac C, Leroux JP, & Charpentier C (1978) Dichloroacetate as treatment for congenital lactic acidosis. N Engl J Med, 299: 1365-1366. Couri D & Abdel-Rahman MS (1980) Effect of chlorine dioxide and metabolites on glutathione dependent system in rat, mouse and chicken blood. J Environ Pathol Toxicol, 3: 451-460. Couri D, Miller CH, Bull RJ, Delphia JM, & Ammar EM (1982a) Assessment of maternal toxicity, embryotoxicity and teratogenic potential of sodium chlorite in Sprague-Dawley rats. Environ Health Perspect, 46: 25-29. Couri D, Abdel-Rahman MS, & Bull RJ (1982b) Toxicological effects of chlorine dioxide, chlorite and chlorate. Environ Health Perspect, 46: 13-17. Cove DJ (1976) Chlorate toxicity in Aspergillus nidulans: Studies of mutants altered in nitrate assimilation. Mol Gen Genet, 146: 147-159. Crabb DW & Harris RA (1979) Mechanism responsible for the hypoglycemic actions of dichloroacetate and 2-chloropropionate. Arch Biochem Biophys, 198: 145-152. Crabb DW, Yount EA, & Harris RA (1981) The metabolic effects of dichloroacetate. Metabolism, 30: 1024-1039. Cragle DL, Shy CM, Struba RJ, & Stiff EJ (1985) A case-control study of colon cancer and water chlorination in North Carolina. In: Jolley RL, Bull RJ, & Davis WP ed. Water chlorination: Chemistry, environmental impact, and health effects. Chelsea, Michigan, Lewis Publishers, Inc., vol 5, pp 153-159. Craun GF (1985) Epidemiologic studies of organic micro-pollutants in drinking water. Sci Total Environ, 47: 461. Craun GF (1991) Epidemiologic studies of organic micropollutants in drinking water. In: Hutzinger O ed. The handbook of environmental chemistry - Volume 5A: Water pollution. Berlin, Heidelberg, New York, Springer-Verlag, p 1-44. Craun GF, Clark RM, Doull J, Grabow W, Marsh GM, Okun DA, Sobsey MD, & Symons JM (1993) In: Craun GF ed. Safety of water disinfection: Balancing chemical and microbial risks - Conference conclusions. Washington, DC, ILSI Press, pp 657-667. Crebelli R, Conti G, Conti L, & Carere A (1985) Mutagenicity of trichloroethylene, trichloroethanol and chloral hydrate in Aspergillus nidulans. Mutat Res, 155: 105-111. Crozes G, White P, & Marshall M (1995) Enhanced coagulation: its effects on NOM removal and chemical costs. J Am Water Works Assoc, 87(1): 78-89. Crump KS, Hoel D, Langley H, & Peto R (1976) Fundamental carcinogenic processes and their implications to low dose risk assessment. Cancer Res, 36: 2973, 2979-2357. Crump KS & Guess HA (1982) Drinking water and cancer: Review of recent epidemiological findings and assessment of risks. Annu Rev Public Health, 3: 339. Cucinell SA, Odessky L, Weiss M, & Dayton PG (1966) The effect of chloral hydrate on bishydroxycoumarin metabolism. J Am Med Assoc, 197: 366-368. Curry SH, Chu P-I, Baumgartner TG, & Stacpoole PW (1985) Plasma concentrations and metabolic effects of intravenous sodium dichloroacetate. Clin Pharmacol Ther, 37: 89-93. Curry SH, Lorenz A, Chu P-I, Limacher M, & Stacpoole PW (1991) Disposition and pharmacodynamics of dichloroacetate (DCA) and oxalate following oral DCA doses. Biopharm Drugs Dispos, 12(5): 375-390. Cunliffe DA (1991) Bactericidal nitrification in chloraminated water supplies. App Environ Microbiol, 57(11): 3399-3402. Dalhamn T (1957) Chlorine dioxide: Toxicity in animal experiments and industrial risks. Arch. Ind Health, 15: 101-107. Daniel FB, Schenck KM, Mattox JK, Lin EL, Haas DL, & Pereira MA (1986) Genotoxic properties of haloacetonitriles: Drinking water by-products of chlorine disinfection. Fundam Appl Toxicol, 6: 447-453. Daniel FB, Robinson M, Condie LW, & York RG (1990a) Ninety-day oral toxicity study of dibromochloromethane in Sprague-Dawley rats. Drug Chem Toxicol, 13: 135-154. Daniel FB, Condie LW, Robinson M, Stober JA, York RG, Olson GR, & Wang S-R (1990b) Comparative 90-day subchronic toxicity studies on three drinking water disinfectants, chlorine, monochloramine and chlorine dioxide, in the Sprague-Dawley rats. J Am Water Works Assoc, 82: 61-69. Daniel FB, Ringhand HP, Robinson M, Stober JA, Olson GR, & Page NP (1991a) Comparative subchronic toxicity of chlorine and monochloramine in the B6C3F1 mouse. J Am Water Works Assoc, 83: 68-75. Daniel FB, Olson GR, & Stober JA (1991b) Induction of gastrointestinal tract nuclear anomalies in B6C3F1 mice by 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone and 3,4-(dichloro)-5-hydroxy-2(5H)-furanone. Environ Mol Mutagen, 17: 32-39. Daniel FB, DeAngelo AB, Stober JA, Olson GR, & Page NP (1992a)