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    INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY



    ENVIRONMENTAL HEALTH CRITERIA 163





    CHLOROFORM







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

    First draft prepared by Dr. J. de Fouw
    National Institute of Public Health and
    Environmental Protection, Bilthoven,
    Netherlands.

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

    World Health Orgnization
    Geneva, 1994


         The International Programme on Chemical Safety (IPCS) is a
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    International Labour Organisation, and the World Health
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    toxicology. Other activities carried out by the IPCS include the
    development of know-how for coping with chemical accidents,
    coordination of laboratory testing and epidemiological studies, and
    promotion of research on the mechanisms of the biological action of
    chemicals.

    WHO Library Cataloguing in Publication Data

    Chloroform.

        (Environmental health criteria ; 163)

        1.Chloroform - adverse effects 
        I.Series

        ISBN 92 4 157163 2        (NLM Classification: QV 81)
        ISSN 0250-863X

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    CONTENTS

         ENVIRONMENTAL HEALTH CRITERIA FOR CHLOROFORM

    1. SUMMARY

    2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYTICAL
         METHODS

         2.1. Identity
         2.2. Physical and chemical properties
         2.3. Conversion factors
         2.4. Analytical methods
                2.4.1. Sampling and analysis in air
                        2.4.1.1   Direct measurement
                        2.4.1.2   Adsorption-liquid desorption
                        2.4.1.3   Adsorption-thermal desorption
                        2.4.1.4   Cold trap-heating
                2.4.2. Sampling and analysis in water
                2.4.3. Sampling and analysis in biological samples
                        2.4.3.1   Blood and tissues
                        2.4.3.2   Urine
                        2.4.3.3   Fish
                2.4.4. Sampling and analysis in soil gas

    3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

         3.1. Natural occurrence
         3.2. Anthropogenic sources
                3.2.1. Production
                        3.2.1.1   Direct production levels and processes
                        3.2.1.2   Indirect production
                        3.2.1.3   Emissions from direct production and
                                  use
                        3.2.1.4   Emissions from indirect production
                3.2.2. Uses

    4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

         4.1. Transport and distribution between media
                4.1.1. Transport
                4.1.2. Distribution
                4.1.3. Removal from the atmosphere
         4.2. Biotic degradation
         4.3. Bioaccumulation

    5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

         5.1. Environmental levels
                5.1.1. Ambient air
                5.1.2. Indoor air
                5.1.3. Water
                        5.1.3.1   Sea water
                        5.1.3.2   Rivers and lakes
                        5.1.3.3   Rain water
                        5.1.3.4   Waste water
                        5.1.3.5   Ground water
                        5.1.3.6   Drinking-water
                5.1.4. Soil
                5.1.5. Foodstuffs
         5.2. General population exposure
                5.2.1. Outdoor air
                5.2.2. Indoor air
                5.2.3. Drinking-water
                5.2.4. Foodstuffs
         5.3. Occupational exposure during manufacture, formulation or
                use

    6. KINETICS IN LABORATORY ANIMALS AND HUMANS

         6.1. Pharmacokinetics
                6.1.1. Absorption
                        6.1.1.1   Oral
                        6.1.1.2   Dermal
                        6.1.1.3   Inhalation
                6.1.2. Distribution
                6.1.3. Elimination and fate
                6.1.4. Physiologically based pharmacokinetic modelling
                        for chloroform
         6.2. Biotransformation and covalent binding of metabolites
         6.3. Human studies
                6.3.1. Uptake
                        6.3.1.1   Oral
                        6.3.1.2   Dermal
                        6.3.1.3   Inhalation
                6.3.2. Distribution
                6.3.3. Elimination
                6.3.4. Biotransformation

    7. EFFECTS ON LABORATORY MAMMALS AND  IN VITRO TEST SYSTEMS

         7.1. Single exposure
                7.1.1. Lethality
                7.1.2. Non-lethal effects
                        7.1.2.1   Oral exposure
                        7.1.2.2   Subcutaneous and intraperitoneal
                                  exposure

                        7.1.2.3   Inhalation exposure
                        7.1.2.4   Dermal exposure
         7.2. Short-term exposure
                7.2.1. Oral exposure
                        7.2.1.1   Mice
                        7.2.1.2   Rats
                7.2.2. Inhalation exposure
         7.3. Long-term exposure
         7.4. Skin and eye irritation
         7.5. Reproductive toxicity, embryotoxicity and teratogenicity
                7.5.1. Reproduction
                7.5.2. Embryotoxicity and teratogenicity
                        7.5.2.1   Oral exposure
                        7.5.2.2   Inhalation exposure
         7.6. Mutagenicity and related end-points
         7.7. Carcinogenicity
                7.7.1. Mice
                7.7.2. Rats
                7.7.3. Dogs
                7.7.4. Studies on initiating-promoting activity
                        7.7.4.1   Mice
                        7.7.4.2   Rats
         7.8.  In vitro studies
         7.9. Factors modifying toxicity; toxicity of metabolites

    8. EFFECTS ON HUMANS

         8.1. Acute non-lethal effects
         8.2. Epidemiology
                8.2.1. Occupational exposure
                8.2.2. General exposure
         8.3. Abuse and addiction

    9. EFFECTS ON OTHER ORGANISMS IN THE
         LABORATORY AND FIELD

         9.1. Freshwater organisms
                9.1.1. Short-term toxicity
                9.1.2. Long-term toxicity
         9.2. Marine organisms

    10. EVALUATION OF HUMAN HEALTH RISKS AND
         EFFECTS ON THE ENVIRONMENT

         10.1. Evaluation of human health risks
                10.1.1. Exposure
                10.1.2. Health effects
                10.1.3. Approaches to risk assessment
                        10.1.3.1  Non-neoplastic effects
                        10.1.3.2  Neoplastic effects
         10.2. Evaluation of effects in the environment

    11. FURTHER RESEARCH

    12. PREVIOUS EVALUATION BY INTERNATIONAL BODIES

    REFERENCES

    RESUME

    RESUMEN

    WHO TASK GROUP ON ENVIRONMENTAL HEALTH
    CRITERIA FOR CHLOROFORM

     Members

    Dr M.W. Anders, Department of Pharmacology, University of Rochester,
         Rochester, New York, USA

    Dr D.Anderson, British Industrial Biological Research Association
         (BIBRA) Toxicology International, Carshalton, Surrey, United
         Kingdom 

    Dr R.J. Bull, Washington State University, College of Pharmacy,
         Pullman, Washington, USA 

    Dr C.D. Carrington, Food and Drug Administration, Washington DC, USA

    Dr M. Crookes, Environment Section, Building Research Establishment,
         Garston, Watford, United Kingdom 

    Dr E. Elovaara, Institute of Occupational Health, Department  of
         Industrial Hygiene and Toxicology, Helsinki, Finland 

    Dr J. de Fouw, Toxicology Advisory Centre, National Institute  of
         Public Health and Environmental Protection (RIVM), Bilthoven,
         the Netherlands  (Rapporteur)

    Dr M.E. Meek, Environmental Health Directorate, Health Protection
         Branch, Health and Welfare, Ottawa, Canada  (Chairperson)

    Dr R. Pegram, Environmental Toxicology Division, Health Effects
         Research Laboratory, US Environmental Protection  Agency,
         Research Triangle Park, North Carolina, USA

    Dr S.A. Soliman, Department of Pesticide Chemistry, College of 
         Agriculture and Veterinary Medicine, King Saud
         University-Al-Qasseem, Bureidah, Saudi Arabia  (Vice-Chairman)

    Dr L. Vittozzi, Istituto Superiore di Sanità, Laboratorio di 
         Tossicologia, Comparata ed Ecotossicologia, Rome, Italy
          (Vice-Chairman)

    Dr P.P. Yao, Institute of Occupational Medicine, Chinese Academy of
         Preventive Medicine, Beijing, China

     Representatives of other Organizations

    Dr B. Butterworth, International Life Sciences Institute, Risk 
         Science Institute, Washington DC, USA

     Secretariat

    Dr B.H. Chen, International Programme on Chemical Safety, World
         Health Organization, Geneva, Switzerland  (Secretary)

    Dr P.G. Jenkins, International Programme on Chemical Safety,  World
         Health Organization, Geneva, Switzerland

    Dr C. Partensky, International Agency for Research on Cancer,  Lyon,
         France

    NOTE TO READERS OF THE CRITERIA MONOGRAPHS

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



                                 *   *   *



         A detailed data profile and a legal file can be obtained from
    the International Register of Potentially Toxic Chemicals, Case
    postale 356, 1219 Châtelaine, Geneva, Switzerland (Telephone No.
    9799111).



                                 *   *   *



         This publication was made possible by grant number 5 U01
    ES02617-15 from the National Institute of Environmental Health
    Sciences, National Institutes of Health, USA, and by financial
    support from the European Commission.

    ENVIRONMENTAL HEALTH CRITERIA FOR CHLOROFORM

         A WHO Task Group on Environmental Health Criteria for
    Chloroform met in Geneva from 15 to 19 November 1993. Dr B.H Chen,
    IPCS, welcomed the participants on behalf of the Director, IPCS, and
    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 and the environment from exposure to
    chloroform.

         The first draft was prepared by Dr J. de Fouw of the National
    Institute of Public Health and Environmental Protection (RIVM),
    Bilthoven, Netherlands. The second draft was also prepared by Dr
    J.de Fouw incorporating comments received following the circulation
    of the first draft to the IPCS Contact Points for Environmental
    Health Criteria monographs. Dr M.E. Meek (Health and Welfare,
    Canada) made a considerable contribution to the preparation of the
    final text.

         Dr B.H. Chen and Dr P.G. Jenkins, both members of the IPCS
    Central Unit, were responsible for the overall scientific content
    and technical editing, respectively.

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

    ABBREVIATIONS

    ALAT      alanine aminotransferase

    ASAT      aspartate aminotransferase

    Brdu      bromodeoxyuridine

    DENA      diethylnitrosamine

    ENU       ethylnitrosourea

    GGTase    gamma-glutamyl transpeptidase

    LI        labelling index

    NOAEL     no-observed-adverse-effect level

    NOEC      no-observed-effect concentration

    NOEL      no-observed-effect level

    NOLC      no-observed-lethal concentration

    PBPK      physiologically based pharmacokinetics

    SCE       sister-chromatid exchange

    SGPT      serum glutamine-pyruvate transaminase

    UDS       unscheduled DNA synthesis

    1.  SUMMARY

         Chloroform is a clear, colourless, volatile liquid with a
    characteristic odour and a burning, sweet taste. It is degraded
    photochemically, is not flammable and is soluble in most organic
    solvents. However, its solubility in water is limited. Phosgene and
    hydrochloric acid may be formed by chemical degradation.

         Chloroform is used in pesticide formulations, as a solvent and
    chemical intermediate. Its use as an anaesthetic and in proprietary
    medicines is banned in some countries. The commercial production
    amounted to 440 000 tonnes in 1987. Significant amounts of
    chloroform are also produced in the chlorination of water and the
    bleaching of paper pulp.

         There are several analytical methods for the analysis of
    chloroform in air, water and biological materials. The majority of
    these methods are based on direct column injection, adsorption on
    activated adsorbent or condensation in a cool trap, then desorption
    or evaporation by solvent extraction or heating and subsequent gas
    chromatographic analysis.

         It is assumed that most chloroform present in water is
    ultimately transferred to air, due to its volatility. Chloroform has
    a residence time in the atmosphere of several months and is removed
    from the atmosphere through chemical transformation. It is resistant
    to biodegradation by aerobic microbial populations of soils and
    aquifers subsisting on endogenous substrates or supplemented with
    acetate. Biodegradation may occur under anaerobic conditions. The
    bioconcentration in freshwater fish is low. Depuration is rapid.

         Based on estimates of mean exposure from various media, the
    general population is exposed to chloroform principally in food,
    drinking-water and indoor air in approximately equivalent amounts.
    The estimated intake from outdoor air is considerably less. The
    total estimated mean intake is approximately 2 µg/kg body weight per
    day. Available data also indicate that water use in homes
    contributes considerably to levels of chloroform in indoor air and
    to total exposure. For some individuals living in dwellings supplied
    with tap water containing relatively high concentrations of
    chloroform, estimated total intakes are up to 10 µg/kg body weight
    per day.

         Chloroform is well absorbed in animals and humans after oral
    administration but the absorption kinetics are dependent upon the
    vehicle of delivery. After inhalation exposure in humans, 60-80% of
    the inhaled quantity is absorbed. The primary factors affecting the
    absorption kinetics of chloroform following inhalation are its
    concentration and species-specific metabolic capacities. It is
    readily absorbed through the skin of humans and animals and

    significant dermal absorption of chloroform from water while
    showering has been demonstrated. Hydration of the skin appears to
    accelerate absorption of chloroform.

         Chloroform distributes throughout the whole body. Highest
    tissue levels are reached in the fat, blood, liver, kidneys, lungs
    and nervous system. Distribution is dependent on exposure route;
    extrahepatic tissues receive a higher dose from inhaled or dermally
    absorbed chloroform than from ingested chloroform. Placental
    transfer of chloroform has been demonstrated in several animal
    species and humans. Chloroform is eliminated primarily as exhaled
    carbon dioxide. Unmetabolized chloroform is retained longer in fat
    than in any other tissue.

         The oxidative biotransformation of chloroform is catalysed by
    cytochrome P-450 to produce trichloromethanol. Loss of HCl from
    trichloromethanol produces phosgene as a reactive intermediate.
    Phosgene may be detoxified by reaction with water to produce carbon
    dioxide or with thiols including glutathione or cysteine to produce
    adducts. The reaction of phosgene with tissue proteins is associated
    with cell damage and death. Little binding of chloroform metabolites
    to DNA is observed. Chloroform also undergoes P-450-catalysed
    reductive biotransformation to produce the dichloromethyl radical,
    which becomes covalently bound to tissue lipids. A role for
    reductive biotransformation in the cytotoxicity of chloroform has
    not been established.

         In animals and humans exposed to chloroform, carbon dioxide and
    unchanged chloroform are eliminated in the expired air. The fraction
    of the dose eliminated as carbon dioxide varies with the dose and
    the species. The rate of biotransformation to carbon dioxide is
    higher in rodent (hamster, mouse, rat) hepatic and renal microsomes
    than in human hepatic and renal microsomes. Also, chloroform is
    biotransformed more rapidly in mouse than in rat renal microsomes.

         The liver is the target organ for acute toxicity in rats and
    several strains of mice. Liver damage is characterized mainly by
    early fatty infiltration and balloon cells, progressing to
    centrilobular necrosis and then massive necrosis. The kidney is the
    target organ in male mice of other more sensitive strains. The
    kidney damage starts with hydropic degeneration and progresses to
    necrosis of the proximal tubules. Significant renal toxicity has not
    been observed in female mice of any strain.

         Acute toxicity varies depending upon the strain, sex and
    vehicle. In mice the oral LD50 values range from 36 to 1366 mg
    chloroform/kg body weight, whereas for rats, they range from 450 to
    2000 mg chloroform/kg body weight. After a single inhalation
    exposure of 4 h, liver toxicity was observed in mice and rats at
    chloroform levels of 490 and 1410 mg/m3, respectively.

         The most universally observed toxic effect of chloroform is
    damage to the liver. The severity of these effects per unit dose
    administered depends on the species, vehicle and the method by which
    the chloroform is administered. The lowest dose at which liver
    damage has been observed is 15 mg/kg body weight per day
    administered to beagle dogs in a toothpaste base over a period of
    7.5 years. Effects at lower doses were not examined. Somewhat higher
    doses are required to produce hepatotoxic effects in other species.
    Although duration of exposure varied in these studies, the
    no-observed-adverse-effect levels ranged between 15 and 125 mg/kg
    body weight per day.

         Effects in the kidney have been observed in male mice of
    sensitive strains and in the F-344 rat. Severe effects have been
    observed in a particularly sensitive strain of male mice at doses as
    low as 36 mg/kg body weight per day.

         Daily 6 h inhalation of chloroform for 7 consecutive days
    induced atrophy of Bowman's glands and new bone growth in the nasal
    turbinates of F-344 rats. The no-observed-effect level (NOEL) for
    these effects was 14.7 mg/m3 (3 ppm). The significance of these
    effects is being further investigated in longer-term studies.

         Chloroform induced hepatic tumours in mice when administered by
    gavage in corn oil at doses in the range of 138 to 477 mg/kg body
    weight per day. However, when similar doses were administered in
    drinking-water, there was no effect of chloroform on the yield of
    hepatic tumours in mice. Moreover, when chloroform was administered
    in drinking-water as a promoter in initiation/promotion studies, it
    actually appeared to inhibit the development of diethylnitrosamine-
    initiated liver tumours in mice. Thus, the vehicle utilized and/or
    the method in which chloroform is administered is an important
    variable in its induction of hepatic tumours in mice.

         Chloroform induced kidney tumours in rats at doses of 90 to 200
    mg/kg body weight per day in corn oil by gavage. However, in this
    species, results were similar when the chemical was administered in
    the drinking-water, indicating that the response is not entirely
    dependent on the vehicle used.

         The carcinogenic effects of chloroform on the liver and kidney
    of rodents appear to be closely related to cytotoxic and cell
    replicative effects observed in the target organs. The effects on
    cell replication were found to parallel the modifications of
    carcinogenic responses to chloroform that were induced by vehicle
    and mode of administration. The weight of the available evidence
    indicates that chloroform has little, if any, capability to induce
    gene mutation or other types of direct damage to DNA. Moreover,
    chloroform does not appear capable of initiating hepatic tumours in
    mice or of inducing unscheduled DNA synthesis  in vivo. On the

    other hand, hepatic tumours can be efficiently promoted by
    chloroform when it is administered in an oil vehicle. Consequently,
    it is likely that, in the case of prolonged administration of
    chloroform, cytotoxicity followed by cell proliferation is the most
    important cause for the development of liver and kidney tumours in
    rodents.

         There are some limited data to suggest that chloroform is toxic
    to the fetus, but only at doses that are maternally toxic.

         In general, chloroform elicits the same symptoms of toxicity in
    humans as in animals. In humans, anaesthesia may result in death due
    to respiratory and cardiac arrhythmias and failure. Renal tubular
    necrosis and renal dysfunction have also been observed in humans.
    The lowest levels at which liver toxicity due to occupational
    exposure to chloroform has been reported are in the range of 80 to
    160 mg/m3 (with an exposure period of less than 4 months) in one
    study and in the range of 10 to 1000 mg/m3 (with exposure periods
    of 1 to 4 years) in another study. The mean lethal oral dose for an
    adult is estimated to be about 45 g, but large interindividual
    differences in susceptibility occur. There is some weight of
    evidence for an association between exposure to disinfection
    by-products in drinking-water and colorectal and bladder cancer in
    some epidemiological studies. However, these studies are compromised
    by inadequate account of potential confounding factors and other
    weaknesses. The evidence for the carcinogenicity of chlorinated
    drinking-water in humans is inadequate. In addition, the
    disinfection by-products cannot be attributed to chloroform  per se.

         Chloroform is toxic to the embryo-larval stages of some
    amphibian and fish species. The lowest reported LC50 is 0.3
    mg/litre for the embryo-larval stages of  Hyla crucifer. Chloroform
    is less toxic to fish and  Daphnia magna. The LC50 values for
    several species of fish are in the range of 18 to 191 mg/litre.
    There is little difference in sensitivity between freshwater and
    marine fish. The lowest reported LC50 for  Daphnia magna is 29
    mg/litre. Chloroform is of low toxicity to algae and other
    microorganisms.

         The Task Group concluded that the available data are sufficient
    to develop a tolerable daily intake (TDI) for non-neoplastic effects
    and risk-specific intakes for carcinogenic effects of chloroform on
    the basis of studies in animal species; the value will serve as
    guidance in the development of exposure limits by appropriate
    authorities. However, it is cautioned that where local circumstances
    require that a choice must be made between meeting microbiological
    limits or limits for disinfection by-products such as chloroform,
    the microbiological quality must always take precedence. Efficient
    disinfection must  never be compromised.

         Based on the study by Heywood et al. (1979) in which slight
    hepatotoxicity (increases in hepatic serum enzymes and fatty cysts)
    was observed in beagle dogs ingesting 15 mg/kg body weight per day
    in toothpaste for 7.5 years, and incorporating an uncertainty factor
    of 1000 (x10 for interspecies variation, x10 for intraspecies
    variation and x10 for use of an effect level rather than a no-effect
    level and a subchronic study), a TDI of 15 µg/kg body weight per day
    is obtained.

         Based on the available mechanistic data, the approach
    considered most appropriate for provision of guidance based on mouse
    liver tumours is division of a no-effect level for cell
    proliferation by an uncertainty factor. Based on the NOEL for
    cytolethality and cell proliferation in B6C3F1 mice of 10 mg/kg
    body weight per day, following administration in corn oil for 3
    weeks in the study of Larson et al. (1994a) and incorporating an
    uncertainty factor of 1000 (x10 for interspecies variation, x10 for
    intraspecies variation and x10 for severity of effect, i.e.
    carcinogenicity, and less-than-chronic study), a TDI of 10 µg/kg
    body weight per day is obtained.

         It is recognized that the kidney tumours in rats may similarly
    be associated with cell lethality and proliferation. However, since
    data on cell proliferation are not available in the strain where
    tumours were observed and identified information on cell
    proliferation and lethality are short-term (one single gavage and
    7-day inhalation exposure), it is considered premature to deviate
    from the default model (i.e. linearized multistage) as a basis for
    estimation of lifetime cancer risk. The total daily intake
    considered to be associated with a 10-5 excess lifetime risk,
    based on the induction of renal tumours (adenomas and
    adenocarcinomas) in male rats in the study by Jorgenson et al.
    (1985), is 8.2 µg/kg body weight per day.

         Levels of chloroform in surface waters are generally low and
    would not be expected to present a hazard to aquatic organisms.
    However, higher levels of chloroform in surface water resulting from
    industrial discharges or spills may be hazardous to the
    embryo-larval stages of some aquatic species.

    2.  IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYTICAL
        METHODS

    2.1  Identity

    Chemical formula:        CHCl3

    Chemical structure:

                                  H
                                  '
                             Cl - C - Cl
                                  '
                                  Cl

    Common name:             chloroform

    Common synonyms:         trichloromethane, methane trichloride,      
                             trichloroform, methyl trichloride,
                             methenyl trichloride

    CAS chemical name:       chloroform 

    CAS registry number:     67-66-3

    RTECS registry number:   FS 9100000

    2.2  Physical and chemical properties

         The most important physical properties of chloroform (IARC,
    1979; Windholz, 1983) are given in Table 1.

         Chloroform is a clear, colourless, very volatile liquid with a
    characteristic odour and a burning sweet taste. It is not flammable;
    however, the substance may be oxidized by strong oxidizing agents
    with the formation of phosgene and chlorine gas. Pure chloroform is
    light-sensitive. Reagent grade chloroform therefore usually contains
    0.75% ethanol as a stabilizer to avoid photochemical transformation
    to phosgene and hydrogen chloride (IARC, 1979; Budavari, 1989). In
    the absence of light this reaction may be catalysed by iron. By the
    application of stabilizers, such as methanol or ethanol, the
    auto-oxidation may be prevented since the phosgene is fixed as
    carbon dioxide dimethyl (or ethyl) ester. Chloroform stabilized with
    0.006% amylenes is now available. This is important for toxicology
    studies to avoid contamination with by-products that might be formed
    by reaction with ethanol. The substance is soluble in most organic
    solvents, such as alcohol, benzene, ether, petroleum ether, carbon
    tetrachloride, oils and carbon disulfide. Its solubility in water is
    limited.

    Table 1. Physical properties of chloroform

                                                                      

    Colour                                       colourless

    Relative molecular mass                      119.38

    Boiling point at 101.3 kPa                   61.3 °C

    Melting point                                -63.2 °C

    Relative density (20 °C)                     1.484

    Refraction index (Nd 20)                     1.4467

    Heat capacity (20 °C)                        0.979 kJ/kg °C

    Critical temperature                         263.4 °C

    Critical pressure                            5.45 MPa

    Critical density                             500 kg/m3

    Auto-ignition temperature                    > 1000 °C

    Solubility of chloroform in water (25 °C)    7.5-9.3 g/litre

    Heat of combustion                           373 kJ/mol

    Evaporation heat at standard
     boiling point                               247 kJ/kg

    Vapour density (101.3 kPa, 0 °C)             4.36 kg/m3

    Vapour pressure (0 °C)                       8.13 kPa

    Vapour pressure (20 °C)                      21.28 kPa

    Stability                                    air- and light-
                                                 sensitive, breaks down  
                                                 to phosgene, HCl and    
                                                 chlorine

    log Kow (octanol/water partition
     coefficient)                                1.97

                                                                      

         Chloroform produces a hydrate, CHCl3.17H2O, which
    decomposes at 1.6 °C and 8 kPa. In contact with water, at normal
    temperatures in the absence of oxygen, chloroform remains stable. It
    is stable at temperatures up to 290 °C. Heating it in the presence
    of a diluted caustic solution leads to the formation of formic acid.

         The pyrolysis of chloroform vapour at temperatures above 450 °C
    produces tetrachloroethane, hydrochloric acid and various
    chlorinated hydrocarbons. In the presence of potassium amalgam or
    hot copper, acetylene is formed. The reaction with primary amines in
    an alkaline environment is known as the isonitrile reaction;
    aromatic hydroxyaldehydes are formed in the presence of phenolates
    (Reimer-Tiemann reaction). In the Friedel-Crafts reaction,
    chloroform and benzene produce triphenyl methane. Chlorination of
    the compound produces tetrachloromethane; bromination of chloroform
    vapour at 225-275 °C produces CCl2Br2 and CClBr3. Chloroform
    reacts with aluminium bromide to form bromoform (CHBr3).
    Fluoroform (CHF3) is produced in the reaction with hydrogen
    fluoride in the presence of a metallic fluoride as a catalyst.
    Iodoform (CHI3) is produced by allowing chloroform to react with
    ethyl iodide in the presence of aluminium chloride. Unstabilized
    chloroform reacts with aluminium, zinc and iron. Chloroform mixed
    with methanolic sodium hydroxide or acetone, in the presence of a
    base, gives a violent reaction.

    2.3  Conversion factors

         1 mg chloroform/m3 air = 0.204 ppm at 25 °C and 101.3 kPa
         (760 mmHg)

         1 ppm = 4.9 mg chloroform/m3 air

    2.4  Analytical methods

         Many analytical methods for the determination of chloroform
    residues in air, water and biological samples have been reported.
    Table 2 summarizes some of the procedures used in the literature for
    sampling and determining chloroform in different media. The
    detection limits are included in Table 2. Although all of these
    methods were developed to detect chloroform at very low levels, some
    of them can be used only in cases where chloroform is present at
    relatively high levels.

         Since chloroform is very volatile, care must be taken while
    sampling and handling samples to prevent any chloroform from being
    lost during such procedures. In this case, accuracy depends very
    much on the repeatability of the method being used. All but one of
    the methods given in Table 2 use gas chromatographic techniques with
    electron capture detection (ECD), flame ionisation detection (FID),
    photo-ionisation detection (PID) or mass spectrometry (MS) for


        Table 2.  Sampling and analysis of chloroform
                                                                                                                                              
    Medium    Sample method               Analytical method   Detection limit   Sample size     Comments                      Reference
                                                                                                                                              

    Air       aspiration velocity of      MIRAN-infrared      300 µg/m3                         can be used only when         Lioy & Lioy
              28 litres/min, trajectory   spectrometer                                          CHCl3 is presented at         (1983)
              of 20 m                                                                           high levels

    Air       direct injection            GC with a           0.5 µg/m3         5 ml injected   method involves the use of    Lasa et al.
                                          coulometric ECD                                       a continuously operating      (1979)
                                                                                                automatic GC monitor

    Air       direct injection,           GC with two         > 0.4 µg/m3       8 ml injected   efficiency followed from      Lillian &
              calibration gas used for    ECDs installed      (estimated)                       signal ratios of the          Singh (1974)
              reliability                 serially                                              two ECDs

    Air       AIRSCAN/PHOTOVAC            GC-PID              0.5 µg/m3         0.05-1 ml       portable machine, suitable    Leveson et
              direct injection                                                                  for field monitoring          al. (1981)

    Air       adsorption on activated     GC-ECD              approximately     1 m3/24 h       in 1984 the draft standard    NNI (1984)
              charcoal, desorption                            0.1 µg/m3                         NVN 2794 needed to be
              with CS2                                                                          tested for usefulness

    Air       adsorption on Porapak-N,    GC-ECD              1 µg/m3           20 litres       advantage of methanol is the  Van Tassel et
              desorption with 1-2 ml                                                            absence of a background       al. (1981)
              methanol                                                                          signal in the ECD

    Air       adsorption on Porapak-N,    GC-ECD              estimated to      0.3-3 litres    confirmation of results by    Russell &
              thermal desorption at                           be 0.05 µg/m3                     use of GC-MS                  Shadoff (1977)
              200 °C

    Air       adsorption on               GC-ECD-FID two      approximately     1-3 litres                                    Heil et al.
              Chromosorb-102, thermal     detectors           0.06 µg/m3                                                      (1979)
              desorption at 150 °C        positioned in
                                          parallel

                                                                                                                                              

    Table 2 (contd)
                                                                                                                                              
    Medium    Sample method               Analytical method   Detection limit   Sample size     Comments                      Reference
                                                                                                                                              

    Air       adsorption on Tenax,        GC-FID              0.08 µg/m3        2 ml injected                                 Kebbekus &
              sample rate 10-15 ml/min,   GC-MS                                                                               Bozzelli (1982)
              thermal desorption and
              cryofocusing

    Air       adsorption on Tenax-GC,     GC-MS               0.2 µg/m3         20 litres                                     Krost et al.
              cooled with liquid                                                                                              (1982)
              nitrogen, thermal
              desorption at 270 °C

    Air       adsorption on activated     GC-FID with         0.15 mg           up to 30        these two types of detection  Morele et
              coal, desorption with       TCEP,               detector          litres can be   appeared to complement        al. (1989)
              CS2, using                  Chromosorbsen       sitivity          sampled         each other
              methylcyclohexane as IS     column

              adsorption on activated     GC-ECD with 5%      2 µg is 
              coal, desorption with       CV17, Chromosorb    minimum 
              ethanol, using              column              quantifiable 
              trichloroethylene as IS                         value

    Air       collection on charcoal,     GC-FID              0.01 mg per       up to 15        suitable for simultaneous     US NIOSH
              desorption with CS2 using                       sample            litres can be   analysis of two or more       (1984)
              n-undecane as IS                                estimated         sampled         substances

    Air       cold trap, heating the      GC-ECD              0.01 µg/m3        30 ml in        air samples were taken        Harsch &
              cold trap                                                         cold trap       in the stratosphere           Cronn (1978)

    Air       injection into cold trap,   GC-MS (SIM)         0.03 µg/m3        100 ml in                                     Cronn &
              heating the cold trap                                             cold trap                                     Harsch (1979)

                                                                                                                                              

    Table 2 (contd)
                                                                                                                                              
    Medium    Sample method               Analytical method   Detection limit   Sample size     Comments                      Reference
                                                                                                                                              

    Air       cold trap after desication  GC-PID-ECD-FID,     0.005 µg/m3       1 litre         during the process the        Rudolph &
              with magnesium              3 detectors                                           column is kept at -103 °C     Jebsen (1983)
              perchlorate, heating the    placed                                                (cryofocusing)
              cold trap to 257 °C         sequentially

    Breath    collection on Tenax GC      GC-MS               0.11 µg/m3                        suitable for quantitative     Pellizzari
              cartridge, thermal                                                                analysis, one sample in       et al.
              desorption                                                                        1.5 h                         (1985b)

    Water     headspace, CH2Br2 was       headspace GC-ECD    0.02 µg/litre     500 µl          suitable for routine          Herzfeld et
              used as IS                                                        injected        analysis over a wide range    al. (1989)
                                                                                                of differently composed 
                                                                                                river waters

    Water     pentane extraction          GC-ECD using        1 µg/litre        100 ml          suitable for routine          Oliver (1983)
                                          2 mm x 4 mm i.d.                      extracted with  measurements in 
                                          column backed with                    10 ml pentane,  drinking-water
                                          Squalane on                           24 litres of
                                          Chromosorb P                          extract used
                                                                                for injection

    Water     liquid-liquid extraction    GC with a Hall      0.10 µg/litre     3 µl injected   suitable for routine          Mehran et al.
              with pentane                electrolyte                                           analyses                      (1984)
                                          conductivity 
                                          detector, 
                                          Tenax-GC column

    Water     direct aqueous injection    GC-ECD with a       0.02 µg/litre     2 µl injected   suitable for analyses of      Grob (1984)
              of sample into GC           fused silica                                          halocarbons in the 0.01-10
                                          capillary column                                      ppb range

                                                                                                                                              

    Table 2 (contd)
                                                                                                                                              
    Medium    Sample method               Analytical method   Detection limit   Sample size     Comments                      Reference
                                                                                                                                              

    Water     direct aqueous injection    GC-ECD with a       0.1 µl/litre      1 µl injected   easy, fast and reliable       Temmerman &
              of sample into GC           methyl-silicone                                       technique for everyday        Quaghebeur
                                          fused silica                                          quality control               (1990)
                                                                                                capillary column

    Aqueous   diethyl ether extraction    GC-MS with a        < 1 µg/litre      200 ml          suitable for water and        Meier et al.
              with 25 µg                  fused silica        and recovery      extracted,      homogenized environmental     (1985)
              p-bromofluorobenzene        capillary column    efficiency of     extract         samples
              as IS                                           0.85              concentrated
                                                                                to 1 ml, 2 µl
                                                                                injected

    Blood     headspace, magnesium        headspace           0.0225 µg/litre   200 µl          suitable for direct           Aggazzotti
              sulfate heptahydrate and    GC-ECD, with        (2.5 times        injected        measurements of CHCl3         et al.
              n-octyl alcohol were        Chromosorb          standard                                                        (1987)
              added to the plasma         W AW column         deviation)

    Blood     passing inert gas over      GC-MS               3 µg/litre        1-10 ml         suitable for quantitative     Pellizzari
              warmed blood sample,                                                              analysis of CHCl3 in          et al.
              collection on Tenax-GC,                                                           blood                         (1985a)
              thermal desorption

    Blood     diethyl ether extraction    GC-MS with a        qualitative (no   1-5 ml,         suitable for identification   Mink et al.
    plasma    (1:1) with 3 different      fused silica        detection limit   extract         of CHCl3 in biological        (1983)
    and       internal standards added    capillary column    was given)        concentrated    samples
    stomach   to the concentrated                                               to 1 ml of
    contents  extract                                                           of which 2µl
                                                                                is injected

                                                                                                                                              

    Table 2 (contd)
                                                                                                                                              
    Medium    Sample method               Analytical method   Detection limit   Sample size     Comments                      Reference
                                                                                                                                              

    Tissue    maceration in water,        GC-MS               6 µg/kg           5 g             suitable for semi-            Pellizzari
              collection on Tenax-GC,                                                           quantitative analysis of      et al.
              thermal desorption                                                                chloroform in tissues         (1985a)

    Urine     pentane extraction          GC-ECD              < 1 µg/litre      2 µl of         convenient and sensitive      Youssefi
                                                                                extract         means for determining         et al.
                                                                                injected        light halogenated             (1978)
                                                                                                hydrocarbons

    Fish      extraction with pentane     GC-ECD with a       1 µg/kg in        2 µl            extraction efficiency of      Baumann
              and isopropanol,            fused silica        fresh             injected        67%                           Ofstad et
              bromotrichloromethane       capillary column    material                                                        al. (1981)
              used as IS

                                                                                                                                              

    Abbreviations:

       ECD = electron capture detector; FID = flame ionisation detector; GC = gas chromatography; IS = internal standard;
       MS = mass spectrometry; PID = photo-ionisation detector; SIM = selected ion monitoring
    

    measuring chloroform residues. Only the first method listed depends
    on the use of a MIRAN-infrared spectrometer. The sensitivity of this
    method is very poor.

    2.4.1  Sampling and analysis in air

         The methods reported in Table 2 for sampling and analysis of
    chloroform levels in air can be grouped into four different
    categories.

    2.4.1.1  Direct measurement

         In this type of procedure, air is aspirated or injected
    directly into the measuring instrument without pretreatment.
    Although these methods are simple, they can be used only when
    chloroform is present in the air at relatively high levels (e.g.,
    urban source areas, see section 5.1.1).

    2.4.1.2  Adsorption-liquid desorption

         Air samples analysed for their chloroform levels are conducted
    through an activated adsorbing agent (e.g., charcoal or Porapak-N).
    The adsorbed chloroform is then desorbed with an appropriate solvent
    (e.g., carbon disulfide or methanol) and subsequently passed through
    the gas chromatograph (GC) for measurement.

    2.4.1.3  Adsorption-thermal desorption

         In this technique, air samples are also passed through an
    activated absorbing agent (e.g., Tenax-GC, Porapak-Q, Porapak-N or
    carbon molecular sieve). The adsorbed chloroform is then thermally
    desorbed and driven into the GC column for determination.

    2.4.1.4  Cold trap-heating

         In this type of procedure, air samples are injected into a cold
    trap (liquid nitrogen or liquid oxygen are used for cooling). The
    trap is then heated while transferring its chloroform content into
    the packed column of a GC for measurement.

    2.4.2  Sampling and analysis in water

         Several methods of sampling and analysing water for chloroform
    content are included in Table 2. In some of these methods, water
    samples are directly injected into a wide bore or fused silica
    capillary column to which an ECD is attached. In some other water
    analysis procedures mentioned in Table 2, the chloroform in the
    water samples is first extracted by means of a non-polar,
    non-halogenated solvent (e.g.,  n-pentane). Samples of the obtained
    extracts are then injected into the GC for determining chloroform.

    In another procedure, referred to as "close-loop-stripping analysis"
    (CLSA), chloroform is removed from the water sample by purging it
    with a large volume of a gas (e.g., nitrogen); the gas is then
    passed through an adsorption tube and subsequently analysed by
    GC-MS. Using this latter method, a million-fold concentration can be
    achieved, so that chloroform can be quantified even at very low
    levels. A headspace GC technique with ECD has also been used for
    measuring chloroform levels in water samples (see Table 2).

    2.4.3  Sampling and analysis in biological samples

    2.4.3.1  Blood and tissues

         Several procedures for determining chloroform in blood and
    tissue samples are presented in Table 2. A headspace GC technique
    has been used for direct measurement of chloroform in plasma
    obtained from subjects exposed to low levels in air (Aggazzotti et
    al., 1987). The second procedure (Kroneld, 1985) depends on
    liquid-liquid extraction of chloroform from blood samples and
    subsequent injection of the extract into a GC system for
    quantification. In the method of Pellizzari et al. (1985a),
    chloroform is evaporated by passing an inert gas over a warmed
    plasma or macerated tissue sample with adsorption of the vapour on a
    Tenax GC column, and is then recovered by thermal desorption and
    analysed by GC-MS.

    2.4.3.2  Urine

         Youssefi et al. (1978) measured chloroform concentration in
    urine using pentane extraction and GC-ECD analysis.

    2.4.3.3  Fish

         The procedure of Baumann Ofstad et al. (1981) for determining
    chloroform in fish samples is based on extraction by  n-pentane and
    subsequent analysis of the extracts by GC/ECD. It has been reported
    that the sensitivity of this method is greatly affected by the fat
    content of the fish samples.

    2.4.4  Sampling and analysis in soil gas

         Kerfoot (1987) determined the level of chloroform in soil gas
    samples in order to use the results as an indication of ground water
    contamination by this pollutant. In the procedure used, a 75-ml soil
    gas sample was drawn from a depth of 1.3 m by means of a sampling
    probe. The chloroform content of the subsample was directly measured
    in the field using an on-site GC-ECD. The detection limit for
    chloroform in soil gas by this method was reported to be 5 parts per
    billion by volume.

    3.  SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

    3.1  Natural occurrence

         Information on the natural occurrence of chloroform has not
    been identified.

    3.2  Anthropogenic sources

    3.2.1  Production

    3.2.1.1  Direct production levels and processes

         Chloroform was prepared, almost simultaneously in 1831, by the
    action of alkali on chloral (Liebig) and by treating bleaching
    powder with ethanol or acetone (Soubeirain) (Hardie, 1964). It is
    currently manufactured in the USA by hydrochlorination of methanol
    or by chlorination of methane. All chloroform production in Japan
    and western Europe is by chlorination of methane (IARC, 1979). It
    can also be manufactured by oxychlorination of methane (ECDIN,
    1992).

         In the years 1984-1987, the worldwide production of chloroform
    increased from 360 to 440 kilotonnes (see Table 3).

    3.2.1.2  Indirect production

         An important contribution to the total emission of chloroform
    is made through its formation from other substances. In particular
    the reaction of chlorine with organic compounds may produce
    substantial quantities of chloroform. With respect to the formation
    of chloroform in the aquatic medium, it may be assumed that the
    quantities produced are ultimately emitted totally to the
    atmosphere.

         The following sources are known to contribute to the formation
    and emission of chloroform:

    *    Paper bleaching with chlorine (US EPA, 1984; Rosenberg et  al.,
         1991).

    *    Chlorination of drinking-water (US EPA, 1984).

    *    Chlorination of swimming pool water (Bätjer et al., 1980). A 
         study on emissions in indoor public swimming pools in Bremen 
         (Germany) revealed that an average of 10 g chloroform may be 
         produced daily.

    *    Chlorination of cooling water. The quantity of chloroform 
         formed depends on a vast range of factors, such as acidity and 
         the concentration of organic materials.

    Table 3. Chloroform production and production capacity expressed in 
    kilotonnes over a period of 15 years (1973-1988)

                                                                      
    Country                          Year       Production    Capacity
                                                                     

    USA                              1975           118           -
                                     1980           160           -
                                     1984           179           -
                                     1985             -         200
                                     1986           191           -
                                     1987           204           -
                                     1988             -         218

    Japan                            1984            46           -
                                     1985             -          55
                                     1987            55           -
                                     1988             -          60

    Italy                            1973            13           -
                                     1988             -          55

    France                           1973            14           -
                                     1987            45           -
                                     1988             -          55

    Federal Republic of Germany      1973            22           -

    Netherlands                      1973             8           -

    Belgium                          1973            15           -

    European Economic Community      1979            80           -
                                     1980            95           -
                                     1982             -         155
                                     1984           130           -
                                     1985             -         160
                                     1987           150           -
                                     1988             -         200

    World                            1984           360           -
                                     1987           440           -
                                     1988             -         500

                                                                      

    From: ECDIN (1992)

    *    Chlorination of waste water.

    *    Exhaust emissions from traffic. The exhaust fumes of vehicles 
         have been demonstrated to contain chloroform; this originates 
         from the decomposition of 1,2-dichloroethane, which is added 
         to petrol as a lead scavenger (US EPA, 1984). Rem et al. (1982) 
         estimated the amount of chloroform to be 1% of the amount of 
         1,2-dichloroethane added.

    *    Decomposition of trichloroethene in the atmosphere. At high 
         concentrations (1 ppm) in the presence of light and NO2, 1% 
         was estimated to be converted (Appleby et al., 1976). US EPA 
         (1984) estimates this emission to be 780 tonnes/year in the 
         USA.

    *    Decomposition of 1,1,1-trichloroethane has also been suggested
         as a source (van der Heijden et al., 1986).

         Appleby et al. (1976) found that, at relatively high
    concentrations (1 ppm), trichloroethene may yield about 1%
    chloroform under the influence of light and NOx. The estimated
    production of chloroform from trichloroethene is, at most, about 3 x
    106 kg/year; in reality the value is likely to be lower.

         A possible source of chloroform (van der Heijden et al., 1986)
    is its production from 1,1,1-trichloroethane via the photolysis of
    the formed chloral. The increase of chloroform levels in the
    southern hemisphere since 1974 (from 3 to 11 ppt), is in accordance
    with the increase in the levels of 1,1,1-trichloroethane during the
    same period (from 25 to 116 ppt).

    3.2.1.3  Emissions from direct production and use

         Almost all of the emissions arise from production, storage,
    transit and use.

         Estimations of emission factors for the production of
    chloroform range from 0.51 kg chloroform/tonne chloroform
    (controlled) to 3.35 kg chloroform/tonne chloroform (uncontrolled)
    (US EPA, 1984). The Federal Office of the Environment (1981)
    published a higher emission factor of 18 kg chloroform/tonne
    chloroform.

         With respect to emissions of chloroform in the production of
    chlorodifluoromethane, emission factors ranging from 0.077-0.33 kg
    chloroform/tonne chlorodifluoromethane (controlled) to 0.59-2.5 kg
    chloroform/tonne chlorodifluoromethane (uncontrolled) have been
    reported (US EPA, 1984). The Federal Office of the Environment
    (1981) reported an emission factor of 8 kg chloroform/tonne
    chlorodifluoromethane.

    3.2.1.4  Emissions from indirect production

         Significant losses of chloroform can also be expected from
    indirect production of chloroform during the chlorination of water
    and paper pulp. Data on the magnitude of such emissions have not
    been identified.

    3.2.2  Uses

         In the period 1980-1987, the use of chloroform increased in the
    USA from 170 to 200 kilotonnes and in the EEC from 90 to 110
    kilotonnes. The use in Japan was 70 kilotonnes in 1987 (ECDIN,
    1992). Chloroform is used in pesticide formulations, in the
    production of other chemicals, and as a solvent. More than 80% of
    the produced chloroform is used for the production of
    chlorodifluoromethane (ECDIN, 1992). This use is likely to decrease
    in the future due to planned phase-out under the Copenhagen
    Amendment to the Montreal Protocol (1992). Chloroform was formally
    used as an anaesthetic (IARC, 1979).

         In many countries the use of chloroform is banned as an
    ingredient (active or inactive) in human drug and cosmetic products
    (US FDA, 1976). However any drug product containing chloroform in
    residual amounts, resulting from its use as a processing solvent in
    manufacture or as a by-product from the synthesis of an ingredient,
    is not considered to contain chloroform as an ingredient (US FDA,
    1976). Chloroform is registered for use in the USA as an
    insecticidal fumigant for stored barley, corn, oats, popcorn, rice,
    rye, sorghum and wheat (US EPA, 1971). 

    4.  ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

    4.1  Transport and distribution between media

    4.1.1  Transport

         Owing to its relatively high volatility, chloroform is
    preferentially transferred from surface water to air. The
    experimental half-life of chloroform in water (1 ppm solution with a
    depth of 6.5 cm at 25 °C) was found to be 18.5 to 25.7 min in a
    volatilization study by Dilling (1977). In the case of ground
    waters, however, exchange with the atmosphere may not take place as
    readily (Uchrin & Mangels, 1986).

    4.1.2  Distribution

    Adsorption - desorption

         Uchrin & Mangels (1986) described the sorptive behaviour of
    chloroform to solids from the Cohansey (90% sand, 8% silt, 2% clay,
    4.4% organic matter) and Potomac-Raritan-Magothy (70.4% sand, 24%
    silt, 5.6% clay, 2.2% organic matter) aquifer systems, located in
    the southern New Jersey coastal plain. The fact that chloroform
    showed a greater tendency to adsorb to the Cohansey material than to
    the Potomac-RM material might be explained by the difference in
    organic matter content. The organic carbon normalized partition
    coefficient Koc was calculated by Uchrin & Mangels (1986) in two
    ways and appeared to be 57.5 or 70.8. These values are in agreement
    with the Koc values of 86.7 and 63.4 obtained for Cohansey and
    Potomac-RM aquifer solids, respectively. Results from the
    consecutive desorption experiments suggest that the sorption
    processes in the systems used are not completely reversible.

    4.1.3  Removal from the atmosphere

         Since no data on the removal rate of chloroform through
    deposition are available, the values are based on estimates and
    calculations. These values, however, differ widely. The estimated
    half-lives range from 92 to 900 years for wet deposition and from 20
    days to 22 years for dry deposition.

         The calculated half-lives for chloroform degradation are
    reported to be approximately 100 to 180 days. Reaction with hydroxyl
    radicals is likely to be the only mechanism for the decomposition of
    chloroform in the atmosphere (van der Heijden et al., 1986). Cox et
    al. (1976) determined the relative rate constant for chloroform in
    comparison with methane in smog chamber studies to be K = 270
    ppm-1 min-1. However, it is known that the decomposition of
    chlorinated hydrocarbons may lead to intermediary products that can
    accelerate the decomposition process. Dimitriades et al. (1983)

    noted that, in a smog chamber, tetrachloroethene is degraded more
    rapidly than might be expected on the basis of the reaction rate
    constant. Another drawback of the method of Cox et al. (1976) is the
    false assumption that the decomposition of hydrocarbons always leads
    to a transformation of two NO molecules for each carbohydrate
    molecule transformed. The absolute rate constants determined by
    Howard Carleton & Evenson (1976) and by Davis et al. (1976) are in
    agreement with each other, and are K(OH) = 170 ± 20 ppm-1
    min-1 and K(OH) = 160 ± 10 ppm-1 min-1, respectively. Based
    on these rate constants of 170 and 160 ppm-1 min-1, a half-life
    of approximately 60 days can be calculated for the decomposition of
    chloroform in the atmosphere, assuming a 12-h daytime average
    hydroxyl radical concentration of 2 x 10-15 mol/litre (Lyman et
    al., 1982).

         When chloroform is irradiated in the presence of chlorine, a
    rapid reaction takes place, resulting in the formation of radicals.
    At later stages the trichloromethyl radical may also be formed from
    the reaction of CHCl3 with the hydroxyl radical. The
    trichloromethyl radical subsequently reacts with oxygen to form the
    trichloromethyl peroxyl radical, which ultimately leads to the
    formation of phosgene (Spence et al., 1976). This is a possible
    mechanism for the formation of phosgene in ambient air from
    chlorination.

    4.2  Biotic degradation

         Strand & Shippert (1986) reported that chloroform is resistant
    to biodegradation by aerobic microbial communities of soils and
    aquifers subsisting on endogenous substrates or supplemented with
    acetate (Wilson et al., 1981; Bouwer & McCarty 1983). Strand &
    Shippert (1986) used Indianola sandy loam to study the oxidation of
    chloroform to carbon dioxide in natural gas-enriched soils. It
    appeared that some chloroform was oxidized by soils that were
    exposed to cylinder air only, but that the rate in natural
    gas-enriched soils was four times higher. Chloroform oxidation rates
    increased with increasing chloroform concentrations up to 5 µg/g
    soil (see Table 4). Chloroform oxidation continued up to 31 days but
    was inhibited by acetylene and higher concentrations of methane,
    indicating that methane-oxidizing bacteria may catalyse chloroform
    oxidation.

         Bouwer et al. (1981) found significant degradation of
    chloroform in seeded cultures, relative to controls, at initial
    concentrations of 16 and 34 µg/litre. At a high initial chloroform
    concentration of 157 µg/litre, degradation was less evident,
    although there was a gradual reduction in chloroform concentration
    relative to the sterile controls. The anaerobic degradation appeared
    to be the result of biological action, although a combination of
    chemical and biological mechanisms is also possible.

    Table 4. Effect of chloroform concentration on chloroform oxidation

                                                                      
    Applied chloroform concentration       Chloroform oxidized
           (µg/g soil)                       (ng/5 g soil)a
                                                                      

            0.02                                2.8 ± 1.3

            0.11                                8.9 ± 7.7

            0.55                                3.2 ± 7.7

            1.09                               11.1 ± 3.6

            5.47                               20.7 ± 9.6

                                                                      

    a Measured during an 8-day incubation in 5 g of aerobic soil
        acclimated to natural gas

         Chloroform can be degraded by reductive dehalogenation under
    anaerobic conditions. It can be reduced by pure cultures of the
    methanogen  Methanobacterium thermoautotrophicum or the
    sulfate-reducing bacterium  Desulfobacterium autotrophicum (Egli et
    al., 1987). In anaerobic sediments, chloroform is probably degraded
    to carbon dioxide via a carbene mechanism (Bouwer & McCarty, 1983).

         Van Beelen & Van Keulen (1990) studied the degradation of
    radiolabelled chloroform under natural conditions in microcosm
    experiments. In these experiments, the degradation was monitored by
    the appearance of radiolabelled carbon dioxide rather than by the
    disappearance of chloroform. This has the advantage that sorption,
    which can also lead to disappearance of chloroform, does not
    interfere with the measurements. At a concentration of 4 µg
    chloroform/litre, the degradation followed first-order kinetics,
    with half-lives of 12 days at 10 °C and 2.6 days at 20 °C. At a
    concentration of 400 µg chloroform/litre, the degradation rate
    increased with time. After 63 days, the final percentages of label
    in carbon dioxide and chloroform happened to be similar to the
    values of the 4-µg/litre experiment. At the other time intervals the
    percentages of formed carbon dioxide were lower at the higher
    chloroform concentration. Evidently the degradation rate of
    chloroform at 400 µg/litre increases with time due to adaptation of
    the bacteria in the sediment.

    4.3  Bioaccumulation

         Anderson & Lusty (1980) determined bioaccumulation in four
    species of fish  (Salmo gairdneri, Lepomis macrochirus, Micropterus
     salmoides and Ictalurus punctatus). The bioaccumulation factor (on
    a fresh weight basis) appeared to be maximal in  Salmo gairdneri
    (approximately 10). Depuration was complete in this species within
    48 h. A similar value of 6 (whole body; fresh weight) in Lepomis
    macrochirus was reported by Veith et al. (1978).

    5.  ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

    5.1  Environmental levels

    5.1.1  Ambient air

         An overview of the concentrations of chloroform measured in
    areas far from anthropogenic sources is presented in Table 5.

        Table 5. Reported concentrations of chloroform in remote areas
             (From: van der Heijden et al., 1986).

                                                                                      
             Northern hemisphere                        Southern hemisphere

    Locality                Year    Level       Locality             Year    Level
                                    (µgm3)                                  (µgm3)
                                                                                      

    Cork, Ireland           1974    0.133       Cape Town            1974    < 0.015

    Pacific Ocean           1976    0.044       South Africa         1977    < 0.015
     (N.W.)

    California              1976    0.085       Pacific Ocean        1981    0.105
                                                 30-40°S, 138-146°E

    California              1977    0.100       South Pole           1981    0.08

    Kansas                  1978    0.08        Australia            1981    0.110

    Marshall Islands        1981    0.130       Samoa                1981    0.110

    Cape Meares, Oregon     1981    0.225       Eastern Pacific      1981    0.055
                                                 0-40°S

    Pt Barrow, Alaska       1981    0.195

    Hawaii                  1981    0.160

    Eastern Pacific         1981    0.105
     0-40°N

                                                                                      
    
         Chloroform levels in urban centres may be elevated in
    comparison with concentrations in remote areas. As in the case of
    other countries, levels in ambient air in remote areas of the USA

    range from 0.1 to 0.25 µg/m3. In urban and source-dominated areas,
    concentrations are 0.3-9.9 µg/m3 and 4.1-110 µg/m3, respectively
    (ATSDR, 1991). The population-weighted mean concentration of
    chloroform at 17 urban sites sampled across Canada in 1989 was 0.2
    µg/m3 (Environment Canada, 1991).

         Su & Goldberg (1976) reported chloroform levels of 1-15 µg/m3
    in Japanese and European cities. Hourly average concentrations of
    chloroform in the Netherlands, determined during 1979-1981, were
    generally 0.15 µg/m3 or less (estimated detection limit), the
    maximum value being 10 µg/m3 (Den Hartog, 1980, 1981). Average
    concentrations of chloroform during 1990 in four German cities
    (Berlin, Tübingen, Freudenstadt and Leipzig) ranged from 0.26 to 0.9
    µg/m3; the maximum value was 30 µg/m3 detected in Tübingen
    (Toxicology and Environmental Health Institute of Munich Technical
    University, 1992).

    5.1.2  Indoor air

         In a study conducted by the US EPA, volatile organic compounds
    including chloroform were determined in breath, breathing zone air,
    fixed outdoor air, drinking-water and some foodstuffs of populations
    in the USA (Wallace, 1987). The observed increase in the median
    concentration of indoor versus outdoor air (approximately 85%) was
    considered to be consistent with assumptions concerning daily water
    use and likely release of chloroform from water into air (Wallace,
    1987). Based on a survey conducted in 1981 in the Federal Republic
    of Germany, Bauer (1981) reported that levels of chloroform may be
    higher in kitchens where foodstuffs and water are heated.

         Taketomo & Grimsrud (1977) reported average indoor air
    concentrations of chloroform to be 0.3 µg/m3 in a family house and
    1.0-3.4 µg/m3 in an apartment in Montana, USA, compared to 0.2
    µg/m3 in outdoor air. In a nationwide survey of 757 randomly
    selected one-family houses in Canada sampled over a 10-month period
    in 1991, the mean level of chloroform in indoor air was 4.1 µg/m3;
    the maximum value was 69 µg/m3 (Otson et al., 1992). Ullrich
    (1982) reported comparable concentrations in indoor air (1-3
    µg/m3) in Germany, although data on outdoor air levels in the
    vicinity were not presented. Taketomo & Grimsrud (1977) reported
    indoor air chloroform concentrations of between 2 and 10 µg/m3 in
    buildings other than residences, e.g., restaurants and shops.

         Higher levels of chloroform occur in the air of enclosed
    swimming pools, resulting from water chlorination with sodium
    hypochlorite and subsequent release to air. Over a period of eleven
    months, the levels of chloroform directly above the water surface in
    indoor public swimming pools in Bremen, Germany, ranged from 10 to
    380 µg/m3, with an average of about 100 µg/m3 (Bätjer et al.,
    1980; Lahl et al., 1981a). Ullrich (1982) reported a similar mean

    value in four public swimming pools in Germany. Chloroform levels in
    the air of enclosed swimming pools are a function of several factors
    such as the degree of ventilation, the level of chlorination, water
    temperature, the degree of mixing at the water surface, and the
    quantity of organic precursors present (Lahl et al., 1981a).

    5.1.3  Water

    5.1.3.1  Sea water

         The maximum concentration of chloroform determined in a survey
    of bay water at 172 locations was 1 µg/litre (Pearson & McConnell,
    1975). Reported levels in the open ocean (east Pacific) and off the
    coast of California were 0.015 µg/litre and 0.009-0.012 µg/litre,
    respectively (Su & Goldberg, 1976).

    5.1.3.2  Rivers and lakes

         Concentrations of chloroform in surface water vary, depending
    upon the proximity to industrial sources. Concentrations of up to
    394 µg/litre have been reported in rivers in highly industrial
    cities (Ewing et al., 1977; Pellizzari et al., 1979). Levels in
    areas not affected heavily by industrial sources ranged from trace
    to 22 µg/litre (Ohio River Valley Water Sanitation Commission, 1980,
    1982). Concentrations in river water in Germany and Switzerland
    ranged from about 0.01 to 30 µg/litre (Reynolds & Harrison, 1982).
    Average concentrations of chloroform detected in 1989 in German
    rivers ranged from 0.131 to 3.17 µg/litre, with a maximum level of
    5.1 µg/litre detected in the River Main (Toxicology and
    Environmental Health Institute of Munich Technical University,
    1992).

    5.1.3.3  Rain water

         Kawamura & Kaplan (1983) measured 0.25 µg chloroform/litre in
    Los Angeles rain water samples taken in the spring of 1982.

    5.1.3.4  Waste water

         Based on two to four samplings at each of 37 plants (22
    branches of industry), Van Luin & Van Starkenburg (1984) detected
    chloroform mainly in the waste water of flavouring and
    pharmaceutical industries at concentrations of 300 and 16 µg/litre,
    respectively. Concentrations were lower in the waste water of
    slaughter-houses, laundries, and textile, rubber and dye industries.
    In waste-water discharges from the treatment of sewage and
    industrial wastes in the USA, chloroform was detected at
    concentrations ranging from 7.1 to 12.1 µg/litre (Europ-Cost, 1976).

    5.1.3.5  Ground water

         Concentrations of chloroform in ground water vary widely,
    depending principally on proximity to hazardous waste sites (ATSDR,
    1993). Chloroform was detected at levels ranging from 11 to 866
    µg/litre in samples from 5 out of 6 monitoring wells drilled 64 m
    apart in a direction perpendicular to the northward flow of ground
    water at a contaminated site in Pittman, Nevada, USA (the depth of
    unconfined ground water was 2 to 4 m at this selected site)
    (Kerfoot, 1987). In a survey of potentially contaminated sites
    conducted by the US EPA, chloroform was detected at 45% of the
    sites. The median and maximum concentrations were 1.5 and 300
    µg/litre, respectively (Westrick et al., 1989). In 8 out of 29 deep
    wells in the Netherlands sampled at least twice since 1980 at
    several depths (± 10 and 25 m below ground level), chloroform was
    detected (limit of detection, 0.1 µg/litre) (Van der Heijden et al.,
    1986).

    5.1.3.6  Drinking-water

         Chloroform can be formed from naturally occurring organic
    compounds during the chlorination of drinking-water with the rate
    and degree of formation being a function primarily of the
    concentrations of chlorine and humic acid, temperature and pH.
    Levels vary seasonally, the concentrations generally being greater
    in summer than winter.

         Stander (1980) detected chloroform in 16 out of 20 tap water
    samples from the USA and western Europe. The highest concentration
    was 60 µg/litre.

         In a national survey of 450 community water supplies in the USA
    sampled in 1978, chloroform was detected in 94% of surface water
    supplies and 34% of ground-water supplies. Median concentrations in
    surface and ground-water supplies were 60 µg/litre and less than the
    detection limit (0.5 µg/litre), respectively (Brass et al., 1981).
    Finished drinking-water collected in 1988 from 35 sources in the
    USA, of which 10 were located in California, sampled in all four
    seasons (spring, summer and autumn in 1988 and winter in 1989),
    contained median concentrations of chloroform ranging from 9.6 to 15
    µg/litre. The overall median for all four seasons was 14 µg/litre
    (Krasner et al., 1989). In a survey conducted in the USA between
    October 1987 and March 1989, the mean concentration in finished
    water for surface water systems serving more than 10 000 people was
    38.9 µg/litre (90th percentile, 74.4 µg/litre). The comparable mean
    value in the distribution system was 58.7 µg/litre (US EPA, 1992).

         In a national survey of the water supplies of 70 communities in
    Canada conducted during the winter of 1976/1977, concentrations of
    chloroform in treated water of the distribution system 0.8 km from

    the treatment plant averaged 22.7 µg/litre (Williams et al., 1980).
    Concentrations at 10 different locations in southern Ontario sampled
    in the early 1980s were 4.5 to 60 µg/litre in water leaving the
    treatment plant and 7.1 to 63 µg/litre one mile from the plant
    (Oliver, 1983).

         Chloroform levels in drinking-water in 100 German cities
    sampled in 1977 ranged from < 0.1 to 14.2 µg/litre and averaged 1.3
    µg/litre. Concentrations were similar in other surveys conducted in
    Germany in the late 1970s and early 1980s (Lahl et al., 1981a).
    Concentrations of chloroform in chlorinated samples of Rhine river
    water were 9 µg/litre, compared to 0.1 µg/litre in untreated water
    from the river (Zoeteman et al., 1982)

         In Japan, chloroform was detected at concentrations of 18 and
    36 µg/litre in drinking-water (Kajino, 1977).

    5.1.4  Soil

         No data on concentrations of chloroform in uncontaminated soil
    have been identified. Chloroform has been detected, however, in 9.9%
    of hazardous waste sites in the USA; the median concentration was
    12.5 µg/kg (ATSDR, 1993).

    5.1.5  Foodstuffs

         Chloroform has been detected in several foodstuffs, in
    particular in decaffeinated coffee (20 µg/kg), olive oil (28 µg/kg),
    pork (10 µg/kg) and sausages (17 µg/kg). Occasionally,
    concentrations were higher: up to 80 µg/kg in coffee and 90 µg/kg in
    sausages. Levels of 1 to 10 µg/kg have been detected in flour
    products, potatoes, cod liver oil, margarine, lard, fish, mussels
    and milk; levels in most foodstuffs, however, were less than 1 µg/kg
    (Bauer, 1981).

         Daft (1988) reported that chloroform was detected in about 90
    of 300 samples in a market-basket survey of 231 "table ready"
    foodstuffs (prepared and cooked as normally served) in the USA, most
    often in fat-containing samples. In a later account, it was reported
    that 2 to 830 µg chloroform/kg food was detected in 68% of 549
    samples of foodstuffs obtained in a market-basket survey, grouped as
    fat, non-fat, grain-based and non-grain-based (average of 71 µg/kg)
    (Daft, 1989).

         Entz et al. (1982) did not detect chloroform in composite
    samples of meat/fish/poultry or in composite samples of oil/fat in
    39 different foods in the USA, although it should be noted that the
    quantification limits were higher (18 to 28 µg/kg) than those in the
    studies described above. However, the authors did detect chloroform
    at a concentration of 17 µg/litre in the composite of dairy foods.

         Concentrations of chloroform in soft drinks range from 3 to 50
    µg/litre, with levels for cola being at the upper end of the range
    (Abdel-Rahman, 1982; Entz et al., 1982; Wallace et al., 1984).

    5.2  General population exposure

         Based on estimates of mean exposure from various media, the
    general population is exposed to chloroform principally in food
    (approximately 1 µg/kg body weight per day), drinking-water
    (approximately 0.5 µg/kg body weight per day) and indoor air (0.3 to
    1 µg/kg body weight per day) in approximately equivalent amounts.
    Estimated intake from outdoor air is considerably less (0.01 µg/kg
    body weight per day). For some individuals living in dwellings
    supplied with tap water containing relatively high concentrations of
    chloroform, exposures may be as high as 10 µg/kg body weight per
    day.

    5.2.1  Outdoor air

         Based on a daily inhalation volume for adults of 22 m3, a
    mean body weight for males and females of 64 kg, the assumption that
    4 out of 24 h are spent outdoors (WHO, in press), and the mean
    levels of chloroform in ambient air in cities presented in section
    5.1.1 (0.2 µg/m3), mean intake of chloroform from ambient air for
    the general population is estimated to be 0.01 µg/kg body weight per
    day.

    5.2.2  Indoor air

         Based on a daily inhalation volume for adults of 22 m3, a
    mean body weight for males and females of 64 kg, the assumption that
    20 out of 24 h are spent indoors (WHO, in press), and the mean
    levels of chloroform in indoor air presented in section 5.1.2 (1 to
    4 µg/m3), mean intake of chloroform from indoor air for the
    general population is estimated to be 0.3 to 1.2 µg/kg body weight
    per day.

         Aggazzotti et al. (1990) determined levels of chloroform in
    samples of plasma of swimmers and visitors taken "a few minutes
    after" exposure at indoor swimming pools with water chloroform
    concentrations of 16.9-47 µg/litre. Concentrations of chloroform in
    the plasma of all 127 subjects who attended the pools averaged 0.82
    µg/litre and ranged from 0.1 to 3 µg/litre, whereas in the plasma
    samples of 40 nonexposed subjects, chloroform was not detected
    (limit of quantification, 0.1 µg/litre). The mean level of
    chloroform in the plasma was significantly higher in swimmers who
    breathed under stress for a long time directly at the surface of the
    water (training for competitions).

         Individuals may be exposed to elevated concentrations of
    chloroform (from chlorinated tap water) during showering (Jo et al.,
    1990a,b).

         After showering for 10 min in water containing 5 to 36 µg
    chloroform/litre, the concentrations of chloroform in the breath of
    six individuals ranged from 6.0 to 21 µg/m3, while none was
    detected (detection limit 0.86 µg/m3) in any of the samples of
    breath collected prior to a shower (Jo et al., 1990b). Based on
    assumptions of an absorption efficiency from the respiratory tract
    of 0.77, a breathing rate of 0.014 m3/min for a 70-kg adult, a
    shower air concentration of 157 µg chloroform/m3 and a ratio of
    body burden resulting from dermal exposure to that of inhalation
    exposure of 0.93, the authors estimated that the average intake of
    chloroform (inhalation and dermal absorption) was 0.5 µg/kg body
    weight per shower for a person weighing 70 kg.

         Based on a review of relevant estimates, Maxwell et al. (1991)
    concluded that the ratio of the dose of chloroform received over a
    lifetime from inhalation to that received from ingestion of
    drinking-water is probably in the range of 0.6-1.5 but could be as
    high as 5.7. The ratio of the dose received dermally compared to
    that received orally over a lifetime from drinking-water was
    considered to be approximately 0.3 but could be as high as 1.8.

    5.2.3  Drinking-water

         Based on a daily volume of ingestion for adults of 1.4 litres
    and a mean body weight for males and females of 64 kg (WHO, in
    press), and the mean levels of chloroform presented in section 5.1.3
    (generally < 20 µg/litre), estimated mean intake of chloroform from
    drinking-water for the general population is less than 0.5 µg/kg
    body weight per day. As discussed by Bauer (1981), actual levels of
    exposure may be less than those estimated on the basis of mean
    levels in drinking-water since most of the chloroform would be
    expelled from drinking-water that is heated before consumption (tea,
    coffee, soups, sauces). For example, approximately 96% of the total
    volatile halogenated hydrocarbon fraction was eliminated in water
    boiling for 5 min, whereas 50-90% was eliminated upon heating at
    70-90 °C (Bauer, 1981). It should be noted, however, that owing to
    the wide variations in concentrations of chloroform in water
    supplies, intake from drinking-water could be considerably greater
    than estimated here for some segments of the general population.

    5.2.4  Foodstuffs

         Based on a daily volume of ingestion of solid foodstuffs for
    reference adults of 1.536 kg and a mean body weight for males and
    females of 64 kg (WHO, in press), and the mean level and percentage
    detection of chloroform in foodstuffs in a market-basket survey

    reported by Daft (1989) (section 5.1.5), estimated daily intake of
    chloroform from foodstuffs is approximately 1 µg/kg body weight per
    day.

    5.3  Occupational exposure during manufacture, formulation or use

         Workers may be exposed to chloroform during, for example, the
    production of chloroform itself, the synthesis of substances derived
    from chloroform (for example chlorodifluoromethane), the use of
    chloroform as a solvent in bleaching of paper, and in sewage
    treatment facilities. Based on a national survey conducted from 1981
    to 1983, NIOSH estimated that approximately 96 000 workers in the
    USA are potentially exposed to chloroform (ATSDR, 1993).

         Chloroform is used as a solvent both industrially and in the
    laboratory; several studies on concentrations in laboratories have
    been published. Taketomo & Grimsrud (1977) reported levels of
    2.3-8.6 mg/m3 in three laboratories in Montana, USA. In an office
    situated in the same building but distant from the laboratories,
    levels were similar; this was attributed to transfer through the
    air-conditioning system. Levels found by NIOSH in laboratories
    ranged from 0.5 to 24.9 mg/m3 (Salisbury, 1982). Time-weighted (4
    h) average levels during laboratory practicals were 0-375 mg/m3
    (Hertlein, 1980).

         Some data on exposure of workers at sewage treatment facilities
    and at indoor pools and spas have also been reported. Lurker et al.
    (1983) reported a maximum level of 0.02 mg/m3 in sewage treatment
    facilities. Maintenance workers, attendants and life guards at
    indoor pools and spas were exposed to 0.025 and 0.075 mg/m3,
    respectively (Armstrong & Golden, 1986; Benoit & Jackson, 1987).

         Generally low levels of chloroform were detected by Rosenberg
    et al. (1991) in a softwood and hardwood kraft pulp mill. Chloroform
    levels ranged from 50 to 290 µg/m3 and from 220 to 5400 µg/m3 in
    the softwood and the hardwood bleaching plants, respectively.

         Chloroform has been and still is often used in dentistry as one
    of the ingredients of root canal sealers or as a solvent. The
    results of a study by Allard & Andersson (1992) showed that a dental
    team could be exposed to quite high concentrations, ranging from 2.2
    to 19.1 mg/m3.

    6.  KINETICS IN LABORATORY ANIMALS AND HUMANS

    6.1  Pharmacokinetics

    6.1.1  Absorption

    6.1.1.1  Oral

         Chloroform is well absorbed after oral administration. After
    intragastric administration of chloroform (75 mg/kg body weight) in
    water or vegetable oil to male Wistar rats, peak blood
    concentrations were observed in about 6 min, but blood
    concentrations were higher (39.3 versus 5.9 µg/ml) with water than
    with olive oil as the vehicle (Withey et al., 1983). The area under
    the blood concentration-time course curve (AUC) after chloroform
    administration in water was 8.7 times greater than the AUC derived
    from vegetable oil delivery.

         Corley et al. (1990) used the data of Withey et al. (1983) to
    compute gavage absorption rate constants, which were 0.6 h-1 and
    5.0 h-1 for corn oil and water, respectively.

    6.1.1.2  Dermal

         Chloroform is absorbed through the intact skin. Most studies
    have examined the systemic appearance of chloroform (or its
    appearance in expired air) to quantify absorption. Tsuruta (1975)
    estimated an absorption rate of 329 nmol/min per cm2 of skin
    surface for pure chloroform in mice, but this study did not correct
    for metabolism. Morgan et al. (1991) measured blood chloroform
    levels in male F-344 rats during 24-h dermal exposures of a shaved
    region of the back to pure chloroform or to aqueous chloroform
    solutions. The blood chloroform level peaked at 51 mg/litre after
    exposure to the pure chemical for 4 to 8 h, and remained about
    constant for the duration of the exposure period. More rapid
    absorption rates were observed during exposure to the aqueous
    solutions, which resulted in peak blood chloroform levels after
    about 2 h. The authors attributed this difference to hydration of
    the skin. Bogen et al. (1992) applied aqueous solutions of
    [14C]-chloroform to most of the body surface of hairless
    guinea-pigs and obtained a permeability coefficient of 0.13 ml/cm2
    per h. This study recovered metabolites as well as expired
    chloroform to measure absorption.

         Indirect evidence of chloroform absorption was obtained by
    observation of damage to kidney tubules in rabbits treated with 1, 2
    or 4 g chloroform/kg applied under an impermeable plastic cuff held
    tightly to the belly of rabbits for 24 h (Torkelson et al., 1976).

    6.1.1.3  Inhalation

         Lehmann & Hasegawa (1910) exposed rabbits to chloroform vapour
    concentrations of around 20, 54 or 80 g/m3. About 35% of the
    inhaled dose was retained during the first hour of the exposure
    period. The fraction retained declined progressively after longer
    periods of exposure (5 to 10% after 4 h; 2% after 8 h). In dogs
    exposed to 73.2 g chloroform/m3, a steady-state blood
    concentration of 354 mg chloroform/litre was reached within 2 h (Von
    Oettingen et al., 1950).

         Corley et al. (1990) developed a pharmacokinetic model for
    chloroform (see section 6.1.4), which was based on inhalation
    studies in a closed-atmosphere chamber (concentrations of 490-24 500
    mg/m3; 100-5000 ppm). Given the same chloroform concentration
    (4900 mg/m3; 1000 ppm), uptake over 6 h in male B6C3F1 mice
    (total body weight = 450 g) was much more rapid and complete than in
    male F-344 rats (total body weight = 690 g). This difference is due
    primarily to the higher rate of chloroform metabolism in mice.

    6.1.2  Distribution

         Cohen & Hood (1969) performed autoradiography studies in male
    NMRI mice after inhalation or intravenous injection of anaesthetic
    doses of chloroform and found high levels of radioactivity in fat
    and liver. Following a 10-min inhalation exposure, the tissue:blood
    ratios at 0, 15 and 120 min post-exposure were 1.56, 2.10 and 6.7
    for the liver and 6.42, 9.25 and 7.18 for fat, respectively. The
    increase in radioactivity in the liver was attributed to the
    accumulation of non-volatile, ether-extractable products. Other
    tissues (blood, brain, muscle, lung and kidney) contained lesser and
    more uniform amounts of radioactivity. Two hours after intravenous
    injection of [14C]-chloroform, non-volitive radioactivity in the
    liver accounted for 2% of the total dose.

         Bergman (1984) studied the distribution of [14C]-chloroform
    in mice after inhalation of 5 µl of [14C]-chloroform (reported
    dose: 280 mg/kg) during 10 min. Whole-body autoradiography,
    immediately after exposure and 2 h thereafter, showed high
    concentrations of radioactivity in fat, blood, lungs, liver,
    kidneys, spinal cord and nerves, meninges and cerebellar cortex.
    After heating and extraction of the sections, it appeared that
    non-volatile radioactivity was bound in the bronchi, nasal mucosa,
    liver, kidneys, salivary glands and in the duodenal contents. High
    levels of volatile or extractable radioactivity were found in
    testes, preputial gland and epididymis.

         Danielsson et al. (1986) observed tissue binding in gestational
    C57BL mice and their fetuses after inhalation of very low
    concentrations of [14C]-chloroform for 10 min, and in 4-day-old

    C57BL mice after intraperitoneal injection of 0.4 µmoles of
    [14C]-chloroform, respectively. The animals were killed 0, 1, 4
    and 24 h after exposure. Low temperature autoradiograms, as well as
    scintillation spectrometry, showed a high uptake of radioactivity
    (volatile and non-volatile) directly after inhalation, especially in
    the respiratory epithelium and liver, fat, lung, brain and segments
    of tubuli in the renal cortex. Tissue-bound (non-volatile)
    radioactivity was found in the respiratory tract, centrilobular
    regions of the liver, salivary glands, and the conjunctiva of the
    eye. Volatile radioactivity was no longer present 24 h after
    exposure and the non-volatile activity had decreased with time in
    all organs measured. Accumulation of non-volatile metabolites was
    also found in the fetal respiratory tract.

         The placental transport of chloroform was first demonstrated by
    Nicloux (1906) in guinea-pigs. Danielsson et al. (1986) reported
    that chloroform was transported to the conceptus at all stages of
    gestation in mice. Non-volatile metabolites of chloroform
    accumulated in the conceptus with time, especially in the amniotic
    fluid at mid-gestation. The fetal uptake of chloroform was low,
    which, according to the authors, was attributable to the low fat
    content in the fetus. An accumulation of non-extractable metabolites
    was found in the fetal respiratory tract in late gestation.

         Withey & Karpinski (1985) exposed Sprague-Dawley rats on the
    17th day of pregnancy to a series of different concentrations of
    chloroform (111 to 1984 ppm; 544 to 9722 mg/m3) for 5 h.
    Chloroform distribution did not appear to be related to fetal
    position in the uterine horn. There was a highly significant
    inter-litter variation in fetal concentration, and additional tests
    showed that the maternal chloroform concentration accounted for only
    part of the variation. However, the fetal and maternal blood
    concentrations were linear functions of the administered dose, with
    a fetal/maternal ratio of 0.316.

         A sex difference in tissue-bound radioactivity in mice given
    [14C]-chloroform was reported by Taylor et al. (1974).
    Autoradiographic studies showed that the renal cortex of male CF/LP,
    CBA and C57BL mice accumulated more label than the renal cortex of
    female mice of the same strains. Treatment with testosterone
    resulted in an increase in tissue binding in the females and
    castration reduced the binding in the males (Taylor et al., 1974).
    Sex differences in renal binding were not found in the rat or monkey
    (Brown et al., 1974b).

    6.1.3  Elimination and fate

         The results of a pharmacokinetic study in male Wistar rats
    indicated that the elimination of chloroform after intravenous
    administration (jugular vein) at dose levels of 3, 6, 9, 12 or 15

    mg/kg body weight followed a three-compartment model.  Chloroform
    was eliminated at a slower rate from fat (half-life of 106 min) than
    from any other tissue examined. The elimination rates from all
    tissues, except fat, were similar to those derived from blood
    analysis (Whithey & Collins, 1980). The elimination half-lives for
    the water and vegetable oil vehicles were 46 and 39 min,
    respectively.

         Various studies on the elimination of chloroform have been
    reported (Paul & Rubinstein, 1963; Van Dyke et al., 1964; Lavigne &
    Marchand, 1974). Corley et al. (1990) exposed B6C3F1 mice and
    Osborne-Mendel rats to a range of chloroform concentrations for 6 h
    and measured the radioactivity in exhaled air, urine, faeces,
    carcass and skin and in the cage wash (Table 6). The fraction of the
    dose exhaled as unchanged chloroform increased with increasing
    exposure concentration in both mice and rats. [14C]-CO2 was the
    major metabolite exhaled. The data indicate partial metabolic
    saturation at the higher doses studied.

         Brown et al. (1974b) administered [14C]-chloroform (60 mg/kg
    body weight) to mice, rats and squirrel monkeys by the oral route.
    The radioactivity was measured in the exhaled air, urine, faeces and
    carcasses up to 48 h after dosing. The recovery percentages (of the
    dose) are listed in Table 7.

         About 50% of the radioactivity in the urine of the mouse and
    the rat consisted of [14C]-urea and [14C]-bicarbonate.
    Auto-radiography revealed biliary excretion of radioactivity in the
    monkey. A high concentration of radioactivity in the bile was
    present as unchanged chloroform.

         The excreted quantities of chloroform and carbon dioxide in the
    rat, as reported by Brown et al. (1974b), correspond to those
    reported by Reynolds et al. (1984), who found that after oral doses
    of 12 or 36 mg chloroform/kg body weight to the rat, about 70% of
    the dose was excreted as carbon dioxide and 12% as chloroform in the
    24 h following oral administration.

    6.1.4  Physiologically based pharmacokinetic modelling for
           chloroform

         Corley et al. (1990) developed a physiologically based
    pharmacokinetic model (PBPK) for mice, rats and humans that
    incorporated literature values for physiological parameters, tissue
    partition coefficients and metabolic constants. The metabolic
    constants were derived from results of rodent  in vivo gas-uptake
    studies and  in vitro metabolic studies with rodent and human (n=9)
    microsomes. The tissue:air partition coefficients were determined by
    a vial-equilibration technique with tissue homogenates.
    Macromolecular binding constants, which define the fraction of the
    total metabolites that bind covalently to proteins, were estimated 

    Table 6. Radioactivity (mg eq/kg body weight) in B6C3F1 mice and
             Osborne-Mendel rats during and up to 48 h after 6-h 
             exposures to [14C]-chloroform (From: Corley et al., 1990)

                                                                           

    Concentration    Exhaled      Exhaled      Urine    Faeces   Residuea
    (ppm)              14C-       14C-CO2
                    chloroform
                                                                            

    Mice

    10                0.03           7.22       0.95      0.05      0.19
    89                0.47          70.35       7.46      1.24      2.32
    366              23.03         217.85      21.24      3.84      9.68

    Rats

    93                0.76          31.84       3.34      0.40      1.09
    356              16.15          54.85       6.53      0.81      2.18
    1041             78.27          89.04      11.83      1.16      3.95

                                                                           

    a Residues comprising total 14C-label present in carcass, skin
      and cage wash at the  end of post-exposure collection period

    Table 7. Percentage recovery of radioactivity after 
             [14C]-chloroform administration 
             (From: Brown et al., 1974b)

                                                                         

    Species              In breath           In faeces     In carcassa
                                             and urine
                 chloroform      CO2
                                                                         

    Mouse          5.2-7.1        84-87       2.1-3.0        1.2-2.3

    Rat               20            67           8             NA

    Monkey            79            18           2             NA

                                                                         

    a NA = not analysed

    from  in vivo binding data obtained following inhalation exposures
    to radiolabelled chloroform. The model parameters that were derived
    for the three species by Corley et al. (1990) are presented in Table
    8.

    Table 8. Parameters used in the physiologically based
             pharmacokinetic model for chloroforma
                                                                 
                            Mouse          Rat        Human
                                                                 

    Partition coefficients

     Blood/air                21.3         20.8        7.43
     Liver/air                19.1         21.1       17.0
     Kidney/air               11.0         11.0       11.0
     Fat/air                 242          203        280
     Rapidly perfused/air     19.1         21.1       17.0
     Slowly perfused/air      13.0         13.9       12.0

    Metabolic and macromolecular binding constants

     VmaxC (mg/h per kg)      22.8          6.8       15.7
     Km (mg/litre)             0.352        0.543      0.448
     fMMBb (h-1), liver        0.003        0.00104    0.00202
     fMMBb (h-1), kidney       0.010        0.0086     0.00931

                                                                 

    a From: Corley et al. (1990)

    b MMB = macromolecular binding of reactive metabolites; 
      fMMB = fraction of MMB of particular organ

         The blood:air partition coefficients for rodents were
    approximately three times greater than for humans. Metabolism was
    described by a single saturable pathway for each species, but in
    mice, equations accounting for enzyme loss had to be incorporated.
    The VmaxC values reflect the greater metabolic capacity of the
    mouse compared to the rat, which has been shown in numerous studies.
    The model generated predictions consistent with experimental data
    for target organ-specific protein binding in rodents as well as
    total chloroform metabolized and total exhaled chloroform in both
    rodents and humans. Predictions of protein binding suggest a
    relative sensitivity ranking for the three species as follows: mouse
    > rat > humans, assuming that equivalent levels of binding produce
    equivalent toxicities in target tissues (Corley et al., 1990).

         Blancato & Chiu (1993) used the PBPK model of Corley et al.
    (1990) to predict the relative contributions of different exposure
    routes to target tissue doses of chloroform in humans. Tissue

    macromolecular binding was predicted as a dose surrogate. With
    respect to liver dose, a 10-min shower was predicted to contribute
    about 25% of the total dose, with 57% from drinking-water. Showering
    was predicted to account for more than 53% of the total dose to the
    kidney, while drinking-water was estimated to contribute only 7% of
    the dose. This difference was attributed to the absence of a
    first-pass effect with dermal absorption and inhalation exposures.

         Gearhart et al. (1993) recently described an additional PBPK
    model for chloroform in B6C3F1 mice. This model accounts for
    decreases in body temperature associated with exposure to high
    chloroform concentrations. The authors contend that the inclusion of
    an enzyme loss equation for mice in the model of Corley et al.
    (1990) was inappropriate and that the incorporation of temperature
    corrections greatly improved the overall fit of gas uptake data. The
    authors also obtained better model simulations of gas-uptake data by
    including a first-order rate constant, which is consistent with  in
     vitro work demonstrating multiple pathways of chloroform
    biotransformation (Pohl, 1979; Testai et al., 1990).

    6.2  Biotransformation and covalent binding of metabolites

         Chloroform may undergo both oxidative and reductive
    biotransformation (Fig. 1). The oxygenation of chloroform is
    catalysed by cytochrome P450 and produces trichloromethanol.
    Elimination of HCl from trichloromethanol gives phosgene as a
    reactive intermediate (Mansuy et al., 1977; Pohl et al., 1977).

         There is considerable evidence available to support this
    reaction mechanism for the formation of phosgene in the
    biotransformation of chloroform: the biotransformation of chloroform
    to phosgene requires NADPH and oxygen. The phosgene formed in the
    biotransformation of chloroform can be trapped by reaction with
    cysteine to give 2-oxothiazolidine-4-carboxylic acid, and the
    biotransformation of [14C]-chloroform in the presence of cysteine
    gives [14C]-2-oxothiazolidine-4-carboxylic acid. When the
    biotransformation of chloroform was studied in the presence of
    [18O]-dioxygen or [35S]-cysteine, [2-18O]- and
    [1-35S]-2-oxothiazolidine-4-carboxylic acid, respectively, are
    formed. Deutero-chloroform is biotransformed more slowly than
    chloroform (Mansuy et al., 1977; Pohl et al., 1977, 1979, 1980; Pohl
    & Krishna, 1978). Moreover, when [36Cl]-chloroform,
    [3H]-chloroform, or [14C]-chloroform were incubated with liver
    microsomes from phenobarbital-treated Sprague-Dawley rats, only
    label from [14C]-chloroform became covalently bound to proteins
    (Pohl et al., 1980).

    FIGURE 1

         Phosgene reacts rapidly with water to give CO2 and HCl as
    products, which explains the formation of CO2 as a metabolite of
    chloroform (Fry et al., 1972; Brown et al., 1974b). Phosgene may
    also react with tissue nucleophiles to form covalently bound
    products (Uehleke & Werner, 1975). Cysteine blocks the covalent
    binding of [14C]-chloroform-derived radioactivity, which supports
    a role for phosgene in the formation of covalent adducts from
    chloroform (Pohl et al., 1977, 1980). Alternatively, phosgene may
    react with glutathione to form S-(chlorocarbonyl)glutathione; this
    intermediate may react with glutathione to give diglutathionyl
    dithiocarbonate (Pohl et al., 1981) or to give glutathione disulfide
    and carbon monoxide as minor products (Ahmed et al., 1977).

         The reductive biotransformation of chloroform is also catalysed
    by cytochromes P450 (Testai & Vittozzi, 1986) (Fig. 1). Reduction of
    chloroform gives rise to the dichloromethyl radical, which has been
    identified by spin trapping and ESR (Tomasi et al., 1985). No
    evidence for the formation of the dichloromethyl carbanion has been
    presented, whereas the formation of chlorocarbene has been ruled out
    (Wolf et al., 1977). The dichloromethyl radical may react
    preferentially with the fatty acid skeleton of phospholipids to give
    covalently bound adducts (De Biasi et al., 1992).

         The balance between the oxidative and reductive
    biotransformation of chloroform depends on several factors,
    including oxygen and chloroform concentrations, animal species,
    strain, enzyme induction, and the site of metabolism. Oxidative
    metabolism is favoured at low (< 0.1 mM) chloroform concentrations
    (Testai et al., 1990, 1991). Under these conditions, the oxygenation
    of chloroform is catalysed by cytochrome P450 2E1 (Brady et al.,
    1989; Guengerich et al., 1991), and covalent binding of chloroform
    metabolites to proteins and lipids in incubation mixtures containing
    mouse (B6C3F1 or C57BL/6J) liver microsomes is higher than in
    incubation mixtures containing rat (Osborne-Mendel or
    Sprague-Dawley) liver microsomes (Testai et al., 1991).

         Chloroform reduction is increased at high substrate
    concentrations (Testai et al., 1990), but oxidative metabolism is
    quantitatively more important. In incubation mixtures containing 5
    mM chloroform, both oxygenation and reduction of chloroform depend
    on the oxygen tension in the incubation flask. Chloroform reduction
    is particularly evident with microsomes from B6C3F1 mice and
    Osborne-Mendel rats. At high chloroform concentrations (approx. 5
    mM), the oxygenation of chloroform may be catalysed by cytochrome
    P450 2B1, as suggested by the induction of the metabolism due to
    pretreatment by phenobarbital (Branchflower et al., 1983; Testai &
    Vittozzi, 1986; Nakajima et al., 1991). Phenobarbital or
    ß-naphthoflavone pretreatment of Sprague-Dawley rats also stimulates
    the formation of reduced intermediates of chloroform (Testai &
    Vittozzi, 1986). Levels of the  in vitro covalent binding of

    [14C]-chloroform metabolites to proteins were higher with hepatic
    microsomes from rabbits and human biopsies than with hepatic
    microsomes from rats or mice (Uehleke & Werner, 1975).

         The  in vitro formation of dichloromethane as a stable
    end-product of chloroform metabolism was addressed in early studies.
    Dichloromethane was detected in mouse liver slices incubated with
    chloroform (Butler, 1961), but not in slices or subcellular
    fractions of rat liver incubated with chloroform (Paul & Rubinstein,
    1963; Rubinstein & Kanics, 1964). These discrepancies, however, may
    have been due to the incubation conditions employed in these early
    studies.  in vivo results with rats, dogs, mice and human
    volunteers exposed to chloroform consistently indicated no
    expiration of dichloromethane (Butler, 1961; Paul & Rubinstein,
    1963; Fry et al., 1972; Brown et al., 1974b).

         Interspecies differences in the oxidative metabolism of
    chloroform have been found  in vivo. After a [14C]-chloroform
    dose of 60 mg/kg body weight, 85%, 66% and 18% was excreted as
    [14C]-CO2 in C57BL, CF/PL and CBA mice, Sprague-Dawley rats, and
    squirrel monkeys, respectively. Expiration of 14C accounted for
    the elimination of most of the remaining dose (recoveries of 93-98%)
    (Brown et al., 1974b). Mink et al. (1986) and Corley et al. (1990)
    also showed that chloroform is metabolized in the mouse to a greater
    extent than in the rat. Corley et al. (1990) demonstrated that the
    covalent binding of [14C]-chloroform metabolites to liver and
    kidney proteins  in vivo was higher in B6C3F1 mice than in
    Osborne-Mendel rats.

         In several strains of mice given [14C]-chloroform, more
    binding occurred in the kidney tissue of males than in that of
    females (Ilett et al., 1973; Taylor et al., 1974). Male DBA mice
    accumulate twice as much radioactivity in their kidneys as do male
    C57BL mice. This strain difference shows intermediate or
    multifactorial heredity (Hill et al., 1975).

         Differences in binding were associated with variations in
    toxicity (Hill et al., 1975; Clemens et al., 1979). The
    nephrotoxicity of chloroform in male mice of susceptible strains
    (see chapter 7) is most probably related to  in situ renal
    metabolic activation of chloroform (Zaleska-Rutczynska & Krus, 1973;
    Hill, 1978; Clemens et al., 1979; Smith & Hook, 1983; Smith et al.,
    1984). Indeed the overall biotransformation of chloroform in both
    sexes is equal, whereas males exhibit more extensive formation of
    renal tissue-bound metabolites than females (Taylor et al., 1974;
    Smith & Hook, 1984). Smith et al. (1985) observed little chloroform
    metabolism in rat (male, Fischer-344) renal cortical microsomes.
    Additional studies, however, have demonstrated chloroform-induced
    cytolethality and regenerative cell damage in male, Fischer-344 rat
    kidney (Larson et al., 1993). Culliford & Hewitt (1957) reported

    that females became more susceptible after pretreatment with
    androgens, and the sensitivity of the males was reduced after
    castration.

         In the rat and mouse, chloroform biotransformation occurs
    mainly in the liver, but other tissues also show metabolic activity.
    After oral administration of chloroform to mice, maximum covalent
    binding in the liver was observed after 3 h, whereas in the kidney,
    maximum binding was found after 6 to 12 h. Binding appears to be
    dose dependent up to doses of 3 mmol/kg body weight. At higher
    doses, a plateau is reached (Ilett et al., 1973). Löfberg & Tjälve
    (1986) studied the extra-hepatic metabolism of [14C]-chloroform in
    Sprague-Dawley rats. Autoradiography was used to localize
    metabolites in freeze-dried, extracted tissues to distinguish
    between total and bound radioactivity.  in vitro autoradiography,
    in which tissue slices were incubated with [14C]-chloroform and
    then examined autoradiographically, showed the capacity of several
    tissues to metabolize [14C]-chloroform: liver, kidney cortex,
    mucosa of the bronchial tree, tracheal mucosa, olfactory and
    respiratory nasal mucosa, Bowman's glands in the olfactory lamina
    propria mucosae, Steno's gland (the lateral nasal gland), mucosa of
    the oesophagus, larynx, tongue, gingiva, cheek, naso-pharyngeal
    duct, pharynx and the soft palate. Furthermore, autoradiographic
    studies showed that a correlation exists between the ability of the
    tissues to retain metabolites  in vivo and the ability of these
    tissues to metabolize chloroform  in vitro.

         The distribution of the covalent binding of 14C to DNA or RNA
    after an intraperitoneal injection of [14C]-chloroform to Balb/c
    mice or to Wistar rats shows several differences from the
    distribution of the covalent binding to tissue proteins (Colacci et
    al., 1991). The highest levels of covalent binding to DNA were
    observed in mouse kidney (3.17 nmol/g DNA) and lung (3.65 nmol/g
    DNA). In mice, binding to liver DNA (0.83 nmol/g DNA) was lower than
    binding to stomach DNA (1.52 nmol/g DNA).

         Differences in DNA binding of chloroform metabolites among rat
    organs have been found to be limited, and the absolute values were
    lower than those seen in mouse organs. In mice, RNA binding levels
    were high in both the liver and kidney; in rats, they were higher in
    the kidney than in other organs. In mice, protein binding was
    highest in the liver (47.5 nmol/g), whereas in rats it was high in
    both the liver (27.4 nmol/g) and kidney (30.7 nmol/g). In
    incubations containing low (< 0.1 mM) chloroform concentrations and
    human liver microsomes, little formation of reactive metabolites was
    seen (Vittozzi et al., 1991). Detectable covalent binding was
    observed in microsomes from some samples of colonic and ileal mucosa
    of human patients but not of male Sprague-Dawley rats (Testai et
    al., 1991).

         Glutathione (GSH) is an important factor controlling the
    binding of chloroform metabolites to proteins and lipids. In  in
     vitro studies, physiological concentrations of GSH (2-5 mM)
    strongly reduced the covalent binding of chloroform metabolites to
    proteins (Sipes et al., 1977; Cresteil et al., 1979; Smith & Hook,
    1984). In later studies, 3 mM GSH blocked the covalent binding of
    chloroform metabolites to proteins and to phospholipid polar heads,
    whereas covalent binding to the phospholipid fatty acyl chains (due
    to the radical metabolite) was only slightly affected (Testai &
    Vittozzi, 1986; Testai et al., 1990, 1991; De Biasi et al., 1992).
    Pretreatment of rats with diethylmaleate or buthionine sulfoximine
    (BSO) increased the binding of administered [14C]-chloroform to
    proteins (Stevens & Anders, 1981). Pretreatment of
    phenobarbital-induced Sprague-Dawley rats with cysteine decreased
    the covalent binding (Stevens & Anders, 1981b). Toxic effects
    paralleled the covalent binding levels after these pretreatments
    (Stevens & Anders, 1981a).

    6.3  Human studies

    6.3.1  Uptake

    6.3.1.1  Oral

         When Fry et al. (1972) dosed eight volunteers with 0.5 g
    chloroform in olive oil in capsules, approximately 50% of the oral
    dose was metabolized to carbon dioxide. Maximal blood levels of 1 to
    5 ng chloroform/litre were achieved after 1.5 h. In two of the
    subjects, the decline in blood levels could be described by a
    two-compartment model with a half-life of 13 min for the initial
    phase and a half-life of 90 min for the second phase. Chiou (1975)
    reanalysed the data obtained from the two subjects mentioned above
    and calculated an apparent volume of distribution of approximately
    160 litres. The author estimated the hepatic first-pass effect to be
    about 32% and the pulmonary first-pass effect to be 16%. Hence after
    a single oral dose of 0.5 g chloroform, about 52% of the dose may be
    available to the system. Pulmonary and metabolic clearances of 0.7
    and 0.6 litre/min, respectively, gave a total body clearance of 1.3
    litre/min.

    6.3.1.2  Dermal

         Jo et al. (1990a) studied the relative contributions of dermal
    and pulmonary uptake of chloroform in individuals taking showers.
    Post-exposure exhaled air concentrations of chloroform were measured
    to estimate chloroform uptake and were 6 to 21 µg/m3 for normal
    showers and 2.4 to 10 µg/m3 for inhalation-only exposure. The
    difference between normal and inhalation-only exposure was
    significant, and the authors concluded that the contribution of
    dermal exposure was approximately equivalent to inhalation exposure.

         Chinery & Gleason (1993) modified an existing PBPK model to
    predict the exhaled air concentration of chloroform in individuals
    exposed to the chemical while showering. Calibration of the model
    with measured exhaled air concentrations of chloroform in
    individuals exposed while showering either with or without dermal
    absorption generated an expected value for skin-blood partitioning
    of 1.2. This assumes a degree of transfer of chloroform from shower
    water into shower air of 61%. The stratum corneum permeability
    coefficient for chloroform was estimated to be within a range of
    0.16-0.36 cm/h, and the expected value was 0.2 cm/h. The estimated
    ratio of dermally:inhaled absorbed doses while showering ranged
    between 0.6 and 2.2 and the expected value was 0.75.

    6.3.1.3  Inhalation

         The inhalation uptake of chloroform in humans was studied by
    Lehmann & Hasegawa in 1910. More recently, Morgan et al. (1970)
    measured the absorption of chloroform after a single inhalation
    exposure to approximately 5 mg of [38Cl]-chloroform. About 80% of
    the chloroform was absorbed under these conditions.

         Prolonged inhalation of anaesthetic concentrations (about 50 g
    chloroform/m3 air) gave rise to blood chloroform concentrations of
    about 100 mg/litre (Smith et al., 1973).

         The relative contribution of inhalation to chloroform uptake
    during showering has been determined (see section 6.3.1.2, Jo et
    al., 1990a).

    6.3.2  Distribution

         Corley et al. (1990) determined partition coefficients for
    human tissues (see Table 8).

         McConnell et al. (1975) analysed chloroform levels in
    postmortem tissue from eight persons (four males and four females,
    48 to 82 years old) living in non-industrial areas of the United
    Kingdom. The chloroform levels (µg/kg wet tissue weight) observed
    were: body fat, 5-68 (mean = 51); liver, 1-10 (mean = 7.2); kidney,
    2-5; and brain, 2-4.

         Phillips & Birchard (1991) reported on a nationwide survey of
    the general population by the US EPA's National Human Adipose Tissue
    Survey. Several hundred fat samples were pooled into 46 composite
    samples by age and geographic region and were analysed. Chloroform
    was detected at levels ranging from 5 to 580 ng/g in 29 of the
    composite samples.

         Dowty et al. (1976) detected chloroform in human maternal and
    placental cord blood. Erickson et al. (1981) found chloroform,

    supposedly originating from environmental exposure, present in
    mother's milk (concentration not specified). Chloroform was
    identified, but not quantified, in mother's milk samples collected
    from 49 lactating women living in the vicinity of chemical
    manufacturing plants or industrial user facilities in Pennsylvania,
    New Jersey and Louisiana, USA (Pellizzari et al., 1982).

    6.3.3  Elimination

         Human volunteers given oral doses of 500 mg chloroform
    eliminated on average 50% of the dose as CO2 and 40% as unchanged
    chloroform during 8 h after dosing (Fry et al., 1972). The amount of
    expired chloroform varied from 18 to 66%, depending on the obesity
    of individuals.

         After administration of 100, 250 or 1000 mg, the authors
    recovered 0, 12 and 65% of the dose in the expired air,
    respectively. After administration of [14C]-chloroform to a man
    and a woman, approximately 50% of the dose was found in the exhaled
    air (as CO2) in 7.5 h after dosing. Virtually no chloroform was
    excreted by the kidneys (Fry et al., 1972).

         After inhalation of chloroform (concentrations of 21 or 35 g
    chloroform/m3), Lehmann & Hasegawa (1910) found little pulmonary
    excretion, i.e. approximately 2% of the absorbed quantity within 30
    min after the exposure. A pulmonary excretion of 10% of the body
    content during the first hour after exposure was reported by Morgan
    et al. (1970).

    6.3.4  Biotransformation

         Human cytochrome P450 2E1 catalyses the oxygenation of
    chloroform (Guengerich et al., 1991).

         Corley et al. (1990) quantified CO2 production from
    incubations of human liver microsomes with 0.049-0.058 mM
    chloroform. The average activity of samples from nine individuals
    was 8.15 ± 0.02 pmol chloroform oxidized/min per mg protein. These
    data were correlated with rodent  in vitro and  in vivo conversion
    rates to estimate human  in vivo metabolic rate constants (see
    Table 8). 

    7.  EFFECTS ON LABORATORY MAMMALS AND  IN VITRO TEST SYSTEMS

    7.1  Single exposure

    7.1.1  Lethality

         The LD50 values of chloroform for mice and rats are given in
    Tables 9 and 10, respectively. Chloroform-induced death is usually
    due to liver damage, with the exception of male mice of very
    sensitive strains, whose death is caused by kidney damage. The
    higher susceptibility to chloroform acute toxicity in these strains
    of mice (such as DBA, C3H, C3Hf, CBA, Balb/c, C3H/He), with respect
    to other strains, is genetically controlled. An absolute sex-related
    difference with respect to kidney damage, but not to liver damage,
    has been described in mice: female mice do not develop renal
    lesions. This is independent of the strain. Some influence of age on
    chloroform acute toxicity in rats has also been described (Kimura et
    al., 1971).

         For the rat, the LD50 values ranged from 450 to 2000 mg
    chloroform/kg body weight, and in this species no sex difference in
    susceptibility was found (Kimura et al., 1971; Chu et al., 1980).
    For OF1 female mice, a LC50 value of 6150 mg chloroform/m3 (6
    h exposure) was reported by Gradiski et al. (1978).

         A dose of 3070 mg chloroform/kg body weight in mineral oil to
    rats resulted in death due to CNS depression within minutes, and a
    dose of 980 mg chloroform/kg body weight resulted in hepatic
    centrilobular necrosis (Reynolds & Yee, 1967). When administered to
    newborn rats, chloroform was lethal at oral doses of 1500 mg/kg body
    weight; smaller doses were not administered (Kimura et al., 1971).

    7.1.2  Non-lethal effects

    7.1.2.1  Oral exposure

         Chloroform is a potent anaesthetic. Anaesthesia may result from
    oral administration of chloroform; this was established by Bowman et
    al. (1978) in the ICR mouse with a dose of 500 mg chloroform/kg body
    weight in aqueous emulsion. The ED50 (50% of animals showing
    effect at this dose level) in mice for acute neurological effects
    (ataxia, incoordination and anaesthesia) was 484 mg chloroform/kg
    body weight (Balster & Borzelleca, 1982).

         After oral administration of chloroform in olive oil to Swiss
    mice (both sexes), Jones et al. (1958) found the median narcotic
    dose to be 350 mg/kg body weight and the median hepatotoxic dose to
    be 35 mg/kg body weight. At this dose level, the liver showed
    centrilobular fatty infiltration and at 350 mg/kg body weight
    centrilobular necrosis was found.

        Table 9.  Representative LD50 values (mg chloroform/kg body weight) for mice
                                                                                               
    Sex/strain    Route             Vehicle        Observed     LD50    Reference
                                                    period
                                                                                               

    Male

    C3H/tif       oral             sesame oil       15 days       36    Pericin & Thomann (1979)
    DBA/2/j       oral             sesame oil       15 days      101    Pericin & Thomann (1979)
    Tif:MAGf      oral             sesame oil       15 days      213    Pericin & Thomann (1979)
    A/J           oral             sesame oil       15 days      253    Pericin & Thomann (1979)
    Tif:MF2f      oral             sesame oil       15 days      336    Pericin & Thomann (1979)
    C57BL/6J      oral             sesame oil       15 days      460    Pericin & Thomann (1979)
    Princeton     subcutaneous     peanut oil       10 days      696    Plaa et al. (1958)
    Swiss albino  subcutaneous     olive oil        10 days     3245    Kutob & Plaa (1962a)

    Female

    C3H/tif       oral             sesame oil       15 days      353    Pericin & Thomann (1979)
    DBA/2/j       oral             sesame oil       15 days      679    Pericin & Thomann (1979)
    A/J           oral             sesame oil       15 days      774    Pericin & Thomann (1979)
    C57BL/6J      oral             sesame oil       15 days      820    Pericin & Thomann (1979)
    Tif:MF2f      oral             sesame oil       15 days     1126    Pericin & Thomann (1979)
    Tif:MAGf      oral             sesame oil       15 days     1366    Pericin & Thomann (1979)
    OF1           intraperitoneal  olive oil        14 days      880    Gradiski et al. (1974)

                                                                                               

    Table 10.  Representative LD50 values (mg chloroform/kg body weight) for rats
                                                                                               
    Sex/strain        Route          Vehicle        Observed     LD50      Reference
                                                    period
                                                                                               

    Male

    Sprague-Dawley    oral             none         unknown       450    Kimura et al. (1971)
     (14 days old)

    unknown           oral             none         unknown      1200    Kimura et al. (1971)
     (older adults)

    Sprague-Dawaley   oral             none         14 days       908    Chu et al. (1980)

    Sprague-Dawley    oral             none         14 days      2000    Torkelson et al. (1976)

    Female

    Sprague-Dawley    oral             none         unknown       450    Kimura et al. (1971)
     (14 days old)

    Sprague-Dawley    oral             none         14 days      1117    Chu et al. (1980)

    Sprague-Dawley    intraperitoneal  peanut oil   24 h         1379    Lundberg et al. (1986)
                                                    14 days       894

                                                                                               
    
         Hill (1978) investigated strain and sex differences in
    chloroform-induced toxicity in mice. Male mice of three strains
    (DBA/2J, B6D2F1/J, and C57BL/6J) were given single oral doses of
    chloroform in oil. No clear difference in hepatotoxicity between
    strains was observed; centrilobular necrosis occurred at doses
    greater than 250 mg/kg body weight in all three strains. In
    contrast, there were differences between species in renal toxicity.
    Doses of 89 mg/kg body weight caused glucosuria and/or proteinuria
    in half of the DBA/2J animals, while doses of 119 and 163 mg/kg body
    weight were required to produce these effects in half the B6D2F1/J
    and C57BL/6J mice, respectively.

         In male CFLP Swiss mice, Moore et al. (1982) found neither
    histological changes in the liver or kidney nor biochemical changes
    in plasma 4 days after oral administration of 17 mg chloroform/kg
    body weight in corn oil. Administration of 66 mg chloroform/kg body
    weight caused slight hepatotoxicity and a more severe
    nephrotoxicity.

         Chu et al. (1980, 1982a) observed piloerection, sedation,
    flaccid muscle tone, ataxia, prostration and dacryorrhoea after
    administration of chloroform to rats. Food intake in the males was
    reduced. Histological and biochemical examination revealed effects
    on liver, kidneys and red and white blood cells. Upon histological
    examination, no lesions were found in other tissues with chloroform
    doses up to 2100 mg/kg body weight. In this study the lowest
    administered dose was 546 mg chloroform/kg body weight, a level at
    which toxic effects were still found.

         Reitz et al. (1982) determined the cellular regeneration (as
    3H-thymidine uptake in DNA) 48 h after administration of
    chloroform to male B6C3F1 mice and male Osborne-Mendel rats. In
    the mice, 3H-thymidine uptake was significantly increased in
    kidneys at a dose level of 60 mg chloroform/kg body weight and in
    kidneys and liver at 240 mg chloroform/kg body weight. In the rats,
    only a slight increase in 3H-thymidine uptake in liver and kidneys
    was found at a dose level of 180 mg chloroform/kg body weight.

         Torkelson et al. (1976) reported dose-related liver and kidney
    changes in adult rats at dose levels as low as 250 mg chloroform/kg
    body weight. Tyson et al. (1983) found an elevation of serum
    aminotransferase levels in rats at dose levels above 200 mg/kg body
    weight in oil.

         One study examined the organ-specific toxicity of acute doses
    of chloroform (Larson et al., 1993). Male F-344 rats were given
    chloroform by gavage in corn oil at doses of 0, 34, 180 or 477 mg/kg
    body weight and necropsied 24 h later. Additional rats were given a
    single dose of 180 mg chloroform/kg and administered
    bromodeoxyuridine (BrdU) 2 h prior to necropsy at 0.5, 1, 2, 4, and
    8 days after chloroform treatment to label cells in S-phase. The
    kidneys of male rats administered 34, 180 and 477 mg chloroform/kg
    exhibited mild to severe proximal tubular necrosis in a
    dose-dependent manner. A 20-fold increase in the labelling index
    (LI, % of nuclei in S-phase) in the proximal tubule cells was
    observed 2 days after treatment at a dose of 180 mg/kg body weight.
    The livers of the male rats exhibited only slight to moderate
    multifocal centrilobular necrosis at 180 and 477 mg/kg body weight.
    A 10-fold increase in the LI was observed in the liver of male rats
    given 477 mg/kg body weight, but no increase was observed at 180
    mg/kg body weight (Larson et al., 1993).

         Female B6C3F1 mice were given chloroform by gavage (0, 34,
    238 or 477 mg/kg body weight) and necropsied 24 h after treatment.
    Additional mice were given a single dose of 350 mg chloroform/kg
    body weight, labelled with BrdU, and necropsied 0.5, 1, 2, 4, and 8
    days after treatment. Female mice developed a dose-dependent
    centrilobular hepatic necrosis at 238 and 477 mg/kg body weight. No
    renal lesions were observed in female mice at any dose. A peak

    increase in LI of 38-fold was observed in hepatocytes in the livers
    of female mice 2 days after treatment with 350 mg chloroform/kg, but
    the increase in LI observed in the kidneys was only 2-fold (Larson
    et al., 1993). These data indicate that acute chloroform-induced
    cytolethality leads to increased cell proliferation and that the
    organ-specific pattern of toxicity is the same as the organ-specific
    pattern of tumour formation (see NCI, 1976a,b, and section 7.7.1).

         Groger & Grey (1979) intubated Colworth Wistar rats (6 of each
    sex per group) daily with chloroform in peanut oil (0 to 50 mg/kg
    body weight) for periods of 1, 5 or 10 days. There were changes in
    the activity of several liver enzymes, the toxicological
    significance of which is unclear.

         Balster & Borzelleca (1982) administered chloroform in water to
    male ICR mice (8-12/group) and examined their performances in a
    battery of neurobehavioural tests (several exposure periods and
    several dose levels). The only effect observed was a reduced
    achievement in an operant behaviour test after dosing with 100 and
    400 mg chloroform/kg body weight in water for 60 days. At the
    chloroform level of 400 mg/kg body weight, about half the treated
    animals died. No adverse effects on behaviour were observed after 90
    days of dosing with 31 mg chloroform/kg body weight in water.

    7.1.2.2  Subcutaneous and intraperitoneal exposure

         A sex difference in toxicity was found after subcutaneous and
    intraperitoneal administration of chloroform to mice. In males, the
    kidney appeared to be more susceptible than in females, in which the
    liver was found to be the target organ. Smith et al. (1983) exposed
    male and female mice of the ICR strain to chloroform doses of 75 to
    1500 mg/kg body weight (subcutaneous and intraperitoneal).
    Hepatotoxicity was dose-related in both sexes from 375 mg
    chloroform/kg body weight upwards. After the subcutaneous
    administration of 375 mg chloroform/kg body weight, an increase in
    the serum alanine aminotransferase (ALAT) and a decrease in the
    liver non-protein sulfhydryl groups (NPSH) were observed.
    Histological examinations showed centrilobular swelling of the liver
    and necrosis of the hepatocytes in both sexes. At 24 h after
    intraperitoneal exposure to 375 mg chloroform/kg body weight, renal
    toxicity was observed in males but not in females. A decrease in the
    renal NPSH concentration of about 60% in males and 20% in females
    was found. The concentration in females, but not in males, returned
    to normal within 24 h post-dosing. Histological examination of male
    kidneys showed proximal tubular lesions with pyknotic nuclei and
    loss of reticular cytoplasmic structure, necrosis of the cells of
    the proximal tubuli and occlusion of the tubular lumens with hyaline
    casts (Smith et al., 1983).

         Skrzypinska et al. (1991) administered chloroform
    intraperitoneally to Balb/c mice as a single dose ranging from 12.5
    to 100% of the approximate lethal dose. At different time periods
    after administration, mice were sacrificed. Serum glutamine-pyruvate
    transaminase (SGPT) and sorbitol dehydrogenase (SDH), as well as
    glutathione (GSH) and malondialdehyde (MDA) levels in the liver,
    were determined. Increased SGPT and SDH levels were found for all
    doses exceeding one eighth of the approximate lethal dose. The
    depletion of GSH level was kept within 40% for all doses. A 2- to
    4-fold increase of hepatic MDA level was found. The depletion of
    hepatic GSH, and to some extent the increase of SGPT and SDH,
    occurred in a biphasic fashion. Dose-effect functions for these
    biochemical alterations could only be constructed for the second
    delayed phase of action. It is postulated that the hepatotoxicity of
    chloroform is mainly dependent on radical formation in the course of
    biotransformation.

         Plaa & Larson (1965) observed renal toxicity after an
    intraperitoneal dose as low as 48 mg chloroform/kg body weight in
    male Swiss mice. The authors reported that chloroform was the most
    potent nephrotoxic agent of 14 short-chain chlorinated hydrocarbons
    in male mice.

         Ahmadizadeh et al. (1984) found an increase in the relative
    kidney weight after intraperitoneal administration of chloroform in
    peanut oil (150 mg/kg body weight) to male DBA mice, but not after
    chloroform administration to DBA female mice or to male or female
    mice of the C57BL strain.

         Hepatic toxicity, which is the predominant effect in most
    species, was found after a parenteral dose of 450 and 150 mg
    chloroform/kg body weight in the rat and the guinea-pig,
    respectively (Klaassen & Plaa, 1969; Divincenzo & Krasavage, 1974).

         Detection of lipoperoxidation in the liver of PB-induced rats
    exposed to chloroform has been reported by several authors (Klaassen
    & Plaa, 1969; Brown, 1972; Brown et al., 1974a; Masuda et al.,
    1980).

         An increased bile duct/pancreatic fluid flow and a changed
    composition of this fluid were observed after an intraperitoneal
    dose of 1500 mg chloroform/kg body weight to rats (Harms et al.,
    1976; Hamada & Peterson, 1977).

         A single, liver-damaging intraperitoneal dose of chloroform led
    to a maximal glutathione depletion in the liver of PB-pretreated
    rats shortly after dosing (1-2 h) but not in saline-treated rats.
    However, the maximal histopathological findings (centrilobular
    necrosis) occurred much later (after about 24 h) (Docks & Krishna,
    1976).

         In dogs, liver toxicity has been found after intraperitoneal
    administration of chloroform. The ED50 for an increased serum ALAT
    activity via this route appears to be 300 mg chloroform/kg body
    weight. At near-ED50 doses, chloroform caused centrilobular
    vacuolization and centrilobular and subcapsular necrosis. The ED50
    for renal dysfunction in the dog appears to be 645 mg chloroform/kg
    body weight (Klaassen & Plaa, 1967).

         Bai et al. (1992) evaluated the suitability of nine different
    serum bile acids (SBA) as markers of chloroform exposure in rats.
    Increases in specific SBA levels were observed following three daily
    intraperitoneal administrations of chloroform at doses as low as 0.1
    mmol/kg body weight. The effects on SBA levels were detectable at
    much lower doses than were effects on histopathological indices or
    on levels of alanine aminotransferase, aspartatetransaminase,
    alkaline phosphatase, bilirubin or total bile acid.

         Chloroform doses as low as 45 mg/kg body weight reduced the
    microsomal Ca++/Mg++-ATP-ase activity (liver microsome calcium
    pump) in rats (Moore, 1980).

    7.1.2.3  Inhalation exposure

         After exposure to chloroform vapour, the same pattern of
    toxicity in mice was observed as after oral, intraperitoneal or
    subcutaneous administration (Deringer et al., 1953; Hewitt, 1956).
    Deringer et al. (1953) found necrosis in the proximal and distal
    convoluted tubules, hyaline casts in the convoluted tubules and
    collecting ducts, calcification of the cortex, and death after
    exposure of male C3H mice to chloroform concentrations of 3400 to
    5400 mg/m3 for 1 to 3 h; anaesthesia was not observed.

         Kylin et al. (1963) exposed female albino mice of an undefined
    strain to chloroform vapour for 4 h and reported hepatotoxic
    effects. At 24 h after exposure to chloroform concentrations of 490
    mg/m3 or more, a concentration-related fatty infiltration was
    observed. From 980 mg chloroform/m3 upwards, necrosis of liver
    cells and a rise in the serum ornithine carbamoyltransferase level
    were seen. In mice, rabbits, guinea-pigs and cats, anaesthesia was
    induced by exposure to chloroform concentrations in the range of 10
    to 100 g/m3 for periods of 30 min to a few hours. In rabbits and
    guinea-pigs such exposures can cause death (review by Lehmann &
    Flury, 1943).

         As with other anaesthetics, prolonged anaesthesia with
    chloroform may result in respiratory depression, cardiac arrhythmia
    and finally in cardiac arrest. Heart failure is probably due to
    increased sensitivity of the heart muscle to adrenaline (Von
    Oettingen et al., 1950; Von Oettingen, 1964). Exposure of rabbits to
    224 mg/m3 for 1 min led to decreased diastolic pressure, reduction

    of the stroke volume, blood pressure and cardiac output, and an
    increase in the peripheral vascular resistance. The cardiac effects
    were probably not due to respiratory effects, as blood oxygen and
    carbon dioxide tension and pH were not significantly changed (Taylor
    et al., 1976).

         Exposure of rats to a chloroform concentration of 49 g/m3 for
    5 h resulted in respiratory acidosis. Liver cells showed swollen
    rough endoplasmic reticulum with a loss of ribosomes, mitochondrial
    lesions, and cistern-like dilatation of tubular areas of the smooth
    endoplasmic reticulum. An accumulation of fat droplets and reduced
    amino acid incorporation into protein were also found in liver cells
    (Scholler, 1966, 1967).

         Brondeau et al. (1983) found increased serum activities of
    glutamate dehydrogenase and sorbitol dehydrogenase after a single 4
    h exposure of male rats to a chloroform concentration of 1410
    mg/m3. The effects were dose-related and at the highest
    concentrations tested (4600 and 5250 mg/m3) serum aspartate
    aminotransferase levels (ASAT) were also increased.

    7.1.2.4  Dermal exposure

         Single application of 1 or 4 g chloroform/kg body weight for 24
    h to the belly of rabbits, under an impermeable plastic cuff,
    resulted in extensive necrosis and weight loss at both levels. The
    kidneys of all animals showed dose-related degenerative changes in
    the tubules. Livers were not grossly affected (see also section 7.4)
    (Torkelson et al., 1976).

    7.2  Short-term exposure

    7.2.1  Oral exposure

    7.2.1.1  Mice

         Condie et al. (1983) dosed male CD1 mice daily with 0, 37, 74
    and 148 mg chloroform/kg body weight in corn oil for 14 days.
    Histological changes turned out to be the most sensitive indicators
    of liver and kidney toxicity. Dose-related effects were observed at
    dose levels from 37 mg/kg body weight upwards. Kidneys showed
    intra-tubular mineralization, epithelial hyperplasia and cytomegaly.
    Livers showed centrilobular cytoplasmic pallor, marked cell
    proliferation and focal inflammation. After 14 days the body weight
    in the highest dose group was reduced.

         Female and male CD1 mice (7-12 animals of each sex per group)
    were administered daily 0, 50, 125 and 250 mg/kg body weight in
    water by gavage for 14 and 90 days (Munson et al., 1982). Many
    histological and biochemical parameters were examined. After 14

    days, the most important effects were a dose-related decrease in the
    number of antibody-forming cells (as IgM response to sheep red blood
    cells) in both sexes (> 50 mg/kg body weight) and an increase in
    the liver weight of males at doses > 125 mg/kg body weight and of
    females at the highest dose level. The serum ASAT level was
    increased in males and females at the highest dose level and serum
    ALAT was increased in females at the highest dose level. After 90
    days, a depression in the number of antibody-forming cells was found
    at the highest dose level in both sexes. In females at the highest
    dose level, a decrease in cell-mediated type hypersensitivity was
    observed. Liver weight was increased after 90 days of exposure to
    doses > 50 mg chloroform per kg body weight in the females and at
    250 mg chloroform/kg body weight in the males. After 90 days of
    exposure, the animals showed a tolerance against a challenging dose
    of 1000 mg chloroform/kg body weight. The kidneys and livers of all
    dosed animals showed histological changes. In the kidneys these
    changes included small intertubular collections of chronic
    inflammatory cells, whereas in the liver they included generalized
    hydropic degeneration of hepatocytes and occasional small focal
    collections of lymphocytes. In females, small amounts of
    extravasated bile were occasionally noted in the sinusoidal Kupffer
    cells.

         Jorgenson & Rushbrook (1980) administered chloroform to female
    B6C3F1 mice for 90 days in the drinking-water at concentrations of
    0, 200, 400, 600, 900, 1800 and 2700 mg/litre (measured daily
    chloroform doses of 0, 34, 66, 92, 132, 263 and 400 mg/kg body
    weight, respectively). In the first week of the experiment some mice
    in the higher dose groups died of dehydration due to reduced
    drinking. Depression of the central nervous system occurred in the
    animals receiving chloroform and was concentration-related. The only
    treatment-related histopathological findings consisted of a mild
    adaptive and transitory fatty change in the livers of animals dosed
    with 66 mg chloroform/kg body weight or more and a mild lymphoid
    atrophy of the spleen at the higher dose levels.

         There is evidence that the vehicle in which chloroform is
    administered significantly affects its toxicity. Bull et al. (1986)
    found that chloroform administered by gavage in corn oil was
    significantly more hepatotoxic than equivalent doses administered in
    an aqueous emulsion (2% Emulphor(R), polyoxyethylated vegetable
    oil, GAF Corp.). Doses of 0, 130 and 270 mg/kg were administered to
    male and female B6C3F1 mice for 90 days. Liver body weight ratios
    were significantly higher in all dose groups and in both sexes when
    chloroform was administered in corn oil. The SGPT level was
    significantly elevated in both sexes at the high dose level of
    chloroform administered in corn oil, but not in those treated with
    the same dose in Emulphor. Mice treated at all levels of chloroform
    in corn oil showed evidence of extensive vacuolation and those
    treated with the high dose in corn oil showed extensive disruption

    of hepatic architecture including cirrhosis. No such pathological
    changes were observed in any of the animals treated with chloroform
    in 2% Emulphor.

         One study contrasted the toxic responses of chloroform
    administered by gavage in corn oil or given  ad libitum in the
    drinking-water (Larson et al., 1994a). Female B6C3F1 mice were
    administered oral doses (0, 3, 10, 34, 90, 238, or 477 mg/kg per
    day) of chloroform dissolved in corn oil for 4 days or for 5
    days/week for 3 weeks, or were continually exposed to chloroform in
    the drinking-water at concentrations of 0, 60, 200, 400, 900 or 1800
    mg/litre for 4 days or 3 weeks, at which time they were necropsied.
    BrdU was delivered via osmotic pumps implanted 3.5 days prior to
    necropsy. Cell proliferation was evaluated as a BrdU labelling index
    (LI) in histological tissue sections. Dose-dependent changes
    included centrilobular necrosis and markedly elevated LI in mice
    given chloroform in corn oil at 238 or 477 mg/kg, the average daily
    doses that produced tumours in the gavage cancer bioassay (NCI,
    1976a,b). The no-observed-effect level (NOEL) for histopathological
    changes was 10 mg/kg body weight per day and for induced cell
    proliferation 34 mg/kg body weight per day. Chloroform given in the
    drinking-water did not increase the hepatic LI after either 4 days
    or 3 weeks in any of the dose groups, nor were any microscopic
    alterations observed in the liver, even though the cumulative daily
    amount of chloroform ingested in the 1800-mg/litre exposure group
    was 329 mg/kg body weight per day (Larson et al., 1994a). Thus, the
    authors concluded that liver detoxification mechanisms are
    overwhelmed when chloroform is given as a single bolus dose, but the
    liver can detoxify the same daily dose if it is given in small
    amounts resulting from sips of water throughout the day. The authors
    also concluded that the sustained increase in LI in the livers of
    mice administered hepatocarcinogenic doses of chloroform in corn
    oil, but not in the case of chloroform in drinking-water, supports
    the hypothesis that chloroform-induced mouse liver cancer is
    secondary to events associated with induced cytolethality and cell
    proliferation (see also NCI, 1976a,b and section 7.7).

    7.2.1.2  Rats

         Chu et al. (1982a) exposed male weanling Sprague-Dawley rats to
    chloroform via drinking-water for 28 days. The following chloroform
    exposure doses were calculated: 0, 0.13, 1.3 and 11 mg/rat per day
    (0, 0.7, 7.4 and 63 mg/kg body weight, respectively). The only
    treatment-related effect observed was a decrease in the neutrophils
    in the 11-mg group. In a 90-day study by Chu et al. (1982b) male and
    female Sprague-Dawley rats were exposed to chloroform via
    drinking-water at dose levels of 0, 0.17, 1.3, 12 and 40 mg/rat per
    day for males and 0, 0.12, 1.3, 9.5 and 29 mg/rat per day for
    females; this was followed by 90 days of recovery. Water and food
    intake were reduced in the highest dose group. At the 40-mg level a

    higher incidence of spontaneous death occurred. Histological
    examination showed mild liver and thyroid lesions, especially in the
    highest dose group. Livers of both males and females showed: an
    increase in cytoplasmic homogeneity; density of the hepatocytes in
    the periportal area; mid-zonal and centrilobular increase in
    cytoplasmic volume; vacuolation due to fatty infiltration and
    occasional nucleic vesiculation; and hyperplasia of biliary
    epithelial cells. Thyroid lesions consisted of a reduction in
    follicular size and colloid density, increase in epithelial cell
    height and occasional collapse of follicles. Liver and thyroid
    lesions diminished in severity during the 90 days recovery period.

         Jorgenson & Rushbrook (1980) administered chloroform in the
    drinking-water to male Osborne-Mendel rats for 90 days at
    concentrations of 0, 200, 400, 600, 900 and 1800 mg/litre
    (calculated to be 0, 20, 38, 57, 81 and 160 mg chloroform/kg body
    weight, respectively). A concentration-related central nervous
    system depression was seen. Body weights in the 160-mg group were
    reduced throughout the study. Biochemical investigations of serum
    showed no important deviations from control values other than a
    dose-related increase in cholesterol at dose levels of 38 mg
    chloroform/kg body weight or more after 60 days and a decrease in
    triglycerides in the highest dose group from 30 days onwards. After
    90 days of administration, however, these parameters were affected
    in the two highest dose groups only. No dose-related
    histopathological changes were reported.

    7.2.2  Inhalation exposure

         The severity of liver injury due to inhaled chloroform is not
    only influenced by the administered concentration but also by the
    shape of the exposure profile. This was observed by Plummer et al.
    (1990), who exposed male black-hooded Wistar rats (36 per group) for
    4 weeks to chloroform vapour as a constant concentration (245
    mg/m3; 24 h/day; 7 days a week) or as repeated concentrations
    (1387 mg/m3; 6 h/day; 5 days a week), with a similar total
    exposure (154 g/m3-hours) for the two ways of exposure (levels
    were monitored). Hepatic injury appeared to be more severe in the
    continuously exposed group, in which microvesicular fatty change was
    the most prominent feature, while focal necrosis was a minor
    feature. Livers of the animals receiving repeated exposures showed
    only minor injuries in the form of scattered hepatocytes containing
    small fat droplets and a few foci of liver cell necrosis.

         Torkelson et al. (1976) exposed male and female rats (10-12 of
    each sex per group), rabbits (2-3 of each sex per group) and
    guinea-pigs (8-12 of each sex per group) to concentrations of 0,
    110, 230 and 410 mg chloroform/m3 air for 7 h/day, 5 days/week,
    during 6 months. In the male and female rats, relative kidney weight
    was increased at all exposure levels. In the males, at all levels,

    kidneys showed cloudy swelling of the tubular epithelium and the
    livers showed lobular granular degeneration with focal necrosis. At
    the higher exposure levels the effects became more pronounced. The
    effects observed in the males exposed to 110 mg chloroform/m3
    disappeared within 6 weeks after exposure. At 410 mg
    chloroform/m3, death, due to interstitial pneumonitis, occurred in
    the males. No effects were seen in the male rats after 1, 2 or 4 h
    of exposure to 110 mg chloroform/m3 (same schedule of exposure).
    The results obtained after exposure of rabbits and guinea-pigs were
    inconsistent because of low numbers of animals and/or the absence of
    dose-effect relationships.

         The toxicity of one-week exposures to inhaled chloroform has
    been investigated in female B6C3F1 mice and in male F-344 rats
    (Larson et al., 1994b; Méry et al., 1994). Rodents were exposed to
    chloroform vapour at concentrations of 0, 4.9, 14.7, 49, 147, 490 or
    1470 mg/m3 (0, 1, 3, 10, 30, 100 or 300 ppm) for 7 consecutive
    days and necropsied on day 8. Cell proliferation was quantified as
    the percentage of cells in S-phase (BrdU labelling index) measured
    by immunohistochemical detection of BrdU-labelled nuclei. Mice
    exposed to 490 or 1470 mg/m3 exhibited centrilobular hepatocyte
    necrosis and severe vacuolar degeneration of mid-zonal and
    periportal hepatocytes, while exposure to 49 or 147 mg/m3 resulted
    in mild to moderate vacuolar changes in centrilobular hepatocytes.
    Slight, dose-related increases in the hepatocyte LI were observed
    for exposure concentrations of 4.9-14.7 mg/m3, while the LI was
    increased more than 30-fold in the 490- and 1470-mg/m3 groups. The
    kidneys of mice were affected only at 1470 mg/m3 exposure, with
    approximately half of the proximal tubules lined by regenerating
    epithelium and an 8-fold increase in the LI of tubule cells compared
    with controls (Larson et al., 1994b).

         In rats, mild centrilobular vacuolation was observed only in
    the livers of animals exposed to 1470 mg/m3. The hepatocyte LI in
    rats was increased only at 490 and 1470 mg/m3 (3-fold and 7-fold
    over control, respectively). The kidneys of the male rats were
    affected only at 1470 mg/m3. About 25 to 50% of the proximal
    tubules were lined by regenerating epithelium in this exposure
    group, while the LI for tubule cells was increased 2-fold over
    controls (Larson et al., 1994b).

         In the nasal passages of rats, chloroform concentrations of 49
    mg/m3 or more induced histopathological changes that exhibited
    clear concentration-related severity. Chloroform-induced changes
    included increased epithelial mucosubstances in the respiratory
    epithelium of the nasopharyngeal meatus, primarily in the rats. A
    complex set of responses was seen in specific regions of the ethmoid
    turbinates, predominantly in the rats. These lesions in the ethmoid
    region, which involved all of the endo- and ectoturbinates, were
    most severe peripherally and generally spared the tissue adjacent to

    the medial airways. These changes were characterized by atrophy of
    Bowman's glands, increased numbers of vimentin-positive cells in the
    periosteum, new bone formation and increased number of periosteal
    cells in S-phase as determined by BrdU incorporation. Additional
    changes were site-specific loss of mucosubstances and loss of
    immunocyto-chemical staining of acini and ducts of Bowman's glands
    for P450-2E1 and pancytokeratin, and loss of P450-2E1 immunostaining
    of the olfactory epithelium. The only change noted in the mice was
    increased periosteal cell proliferation without new bone growth
    (Méry et al., 1994).

    7.3  Long-term exposure

         In a carcinogenicity bioassay, female B6C3F1 mice were
    exposed to 0, 200, 400, 900 or 1800 mg chloroform/litre
    drinking-water (number of animals: 430, 430, 150, 50 and 50,
    respectively) for a period of two years (Jorgenson et al., 1982)
    (see also section 7.7.1). These concentrations (monitored by
    analysis) correspond to time-weighted average daily chloroform doses
    of 0, 34, 65, 130 and 263 mg/kg body weight (Jorgenson et al.,
    1985). Matched controls (50 females) received an amount of water
    without chloroform equal to that consumed by the 1800-mg/litre
    group. Additional mice were used for intermediate biochemical and
    histopathological examination. Early mortality in the high-dose
    group was observed. After 3 months, livers of animals exposed to
    chloroform concentrations of 65 mg/kg body weight or more showed a
    higher fat content than those of the controls (as examined by
    chemical techniques). After 6 months, liver fat content was
    increased in all exposed groups. Data on organ weights were not
    provided.

         In a carcinogenicity bioassay, male Osborne-Mendel rats were
    exposed to 0, 200, 400, 900 or 1800 mg chloroform/litre
    drinking-water (number of animals: 330, 330, 150, 50 and 50,
    respectively) for a period of two years (Jorgenson et al., 1982)
    (see section 7.7.2). These concentrations (monitored by analysis)
    correspond to time-weighted average daily chloroform doses of 0, 19,
    38, 81 and 160 mg/kg body weight (Jorgenson et al., 1985). Matched
    controls received an amount of water without chloroform equal to
    that consumed by the 1800-mg/litre group. Additional rats were used
    for intermediate biochemical and histopathological examination. The
    survival was indirectly proportional to the dose levels.
    Concentration-related decreases in water uptake and growth were
    seen. The latter effects were also observed in the matched controls,
    and thus may be attributed to the reduced intake of water.
    Biochemical examination of blood after 6, 12 and 18 months showed
    that chlorine, potassium, total iron and albumin levels and the
    albumin/globulin ratio tended to be increased after chloroform
    treatment, whereas levels of cholesterol, triglycerides and lactate
    dehydrogenase were decreased in all treated groups. These deviations

    were also observed in the matched controls, but the decreases in
    serum triglycerides and cholesterol levels were more severe at the
    two highest dose levels than in the matched control group. Data on
    organ weights were not provided.

         Beagle dogs were given chloroform in a toothpaste base in
    gelatin capsules, 6 days/week for 7.5 years (Heywood et al., 1979).
    The doses were 15 and 30 mg/kg and there were eight male and eight
    female dogs in each dose group. Dogs given the high dose began to
    show significant increases in SGPT levels at 6 weeks of treatment.
    At the low dose level, significant increases were observed at 34
    weeks and after. Similar effects were not observed in the vehicle
    control (16 dogs of each sex) or untreated control (eight dogs of
    each sex) groups. "Fatty cysts" of the liver were observed in both
    dose groups at the end of this study (see section 7.7.3).

    7.4  Skin and eye irritation

         Adequate data on the skin irritation potential of chloroform
    has not been identified. Torkelson et al. (1976) applied liquid
    chloroform to the rabbit ear and found slight hyperaemia and
    exfoliation after one to four treatments (period between application
    and observation not specified). More frequent application did not
    increase the severity of the injuries. A 24-h application of
    chloroform on a cotton pad on the belly of rabbits produced slight
    hyperaemia, moderate necrosis and eschar formation. Chloroform
    delayed healing of mechanically damaged skin on the application
    site.

         Application of chloroform droplets in the rabbit eye caused a
    transient slight irritation of the conjunctiva and corneal injury. A
    purulent exudate occurred for 2 or more days after the treatment
    (Torkelson et al., 1976).

         Duprat et al. (1976) applied undiluted chloroform into the eyes
    of six New Zealand white rabbits. It produced severe eye irritation,
    with mydriasis and keratitis in all rabbits. Translucent zones in
    the cornea were observed in four animals and a purulent haemorrhagic
    discharge was also reported (number of rabbits unknown). The effects
    had disappeared 2-3 weeks after application, except for one rabbit
    that still showed corneal opacity after 3 weeks.

    7.5  Reproductive toxicity, embryotoxicity and teratogenicity

    7.5.1  Reproduction

         Borzelleca & Carchman (1982) studied the reproductive toxicity
    of chloroform in a three-generation experiment with ICR mice. They
    administered the chemical (0.1% Emulphor in deionised water) via
    drinking-water (in closed bottles) to males (10/group) and females

    (30/group), at concentrations of 0, 0.1, 1 and 5 mg/ml, from 5 weeks
    before F0 mating throughout the entire study until sacrifice of
    the F2b pups. Death occurred among the males and females of the
    highest dose group, and body weights in this group were reduced. At
    1 mg/ml, the body weights of F1b females were also reduced.
    Dose-related hepatotoxicity was found in the F0 and F1b animals
    (symptoms varying from "slight yellow-grey colouring" in the lowest
    dose group to "grey to black discolouration" with large nodules
    (> 3 mm) upon and within the liver in the highest dose group).
    The treatment resulted in reduced fertility, litter size, gestation
    index and viability index in all F1 and F2 generations,
    statistically significant at 5 mg/ml. No evidence for a teratogenic
    potential was obtained.

    7.5.2  Embryotoxicity and teratogenicity

         Chloroform has not been found to be teratogenic but has been
    shown to induce fetotoxic effects.

    7.5.2.1  Oral exposure

         No evidence for a teratogenic effect of chloroform was obtained
    in a three-generation study with ICR mice (Borzelleca & Carchman,
    1982).

         In a study by Thompson et al. (1974), female Sprague-Dawley
    rats (25/group) were intubated with chloroform in corn oil (0, 10,
    25 and 63 mg/kg body weight) twice daily on days 6-15 of gestation.
    A reduced body weight gain and anorexia were seen in the dams of the
    two higher dose groups. Tissues from two dams of each dose group
    were microscopically examined and fatty changes were observed in the
    livers of both females at 63 mg/kg and in one female at 25 mg/kg.
    Other signs of maternal toxicity were not found at these dose
    levels. The fetuses of the 63-mg/kg groups had a smaller weight at
    delivery than those of the control group. The incidence of bilateral
    extralumbar ribs was significantly increased among the fetal
    population of the 63-mg/kg dose group. Other minor visceral and
    skeletal abnormalities were seen, but not at significantly elevated
    levels. In the same study female Dutch-Belted rabbits (15/group)
    were dosed orally with chloroform in corn oil (0, 20, 35 and 50
    mg/kg body weight) once daily during days 6-18 of gestation.
    Administration of chloroform produced a decrease in fetal body
    weight and incomplete ossification of skeletal elements (skull
    bones) in the 20- and 50-mg/kg dose groups. At the highest dose
    level the dams showed decreased weight gain. Signs of embryotoxicity
    and teratogenicity were not observed.

         Ruddick et al. (1983) gave pregnant Sprague-Dawley rats
    (15/group) chloroform in corn oil (0, 100, 200 and 400 mg
    chloroform/kg body weight) daily by gavage from days 6 to 15 of

    gestation. All doses caused reduced weight gain in the dams and
    increased liver weight. At the highest dose level, there was an
    increase in the kidney weight of the dams. Haematological
    examinations showed dose-dependent reductions in haemoglobin and
    haematocrit (14% maximally, both parameters). In the highest dose
    group, the red blood cell count was also reduced.

         According to Burkhalter & Balster (1979), oral administration
    of chloroform to mice from 3 weeks before mating until the end of
    the lactating period (in both sexes the dose was 31 mg/kg body
    weight) did not result in retardation of the development of
    responses to a battery of neurobehavioural tests in the pups.

    7.5.2.2  Inhalation exposure

         Schwetz et al. (1974) reported effects on pregnancy and on the
    incidence of fetal malformations in Sprague-Dawley rats exposed to
    chloroform concentrations of 147, 490 and 1470 mg/m3 (30, 100 and
    300 ppm) for 7 h/day during days 6-15 of gestation (analysis 3 times
    daily showed concentrations of 147, 466 and 1426 mg/m3; 30, 95 and
    291 ppm, respectively). The two highest concentrations were toxic to
    the dams (anorexia and reduced weight gain, increases in relative
    and absolute liver weight). There was a dose-dependent decrease in
    the pregnancy percentage (100% in the control group versus 15% in
    the 1426-mg/m3 group) and in the number of living fetuses per
    litter. An increase was observed in the percentage of
    post-implantation losses (resorptions) in the highest dose group,
    and a dose-dependent increase was seen in the percentage of litters
    with resorptions (from 57% in the control group to 100% at the
    highest concentration). At all exposure levels, fetuses showed
    growth retardation and minor skeletal aberrations (delayed
    ossification of skull and sternebrae). Exposure to 147 mg/m3
    caused minor embryo- and fetotoxicity, and concentrations of 466 and
    1426 mg/m3 in the inhaled air were embryo- and fetotoxic to the
    rat. At the higher levels, subcutaneous oedema and other unspecified
    fetal soft tissue anomalies were also observed.

         Murray et al. (1979) exposed pregnant CF1 mice (35-40/group)
    to 0 and 490 mg/m3 (0 and 100 ppm) for 7 h each day throughout
    days 1-7, 6-15 or 8-15 of gestation. The ability of females to
    maintain pregnancy was significantly decreased after exposure to
    chloroform during days 1-7 or 6-15 of gestation (44 and 43% in the
    treated groups versus 74 and 91% in the respective control groups).
    The dosed animals consumed slightly less food than the control
    animals, resulting in reduced body weight gain. Absolute and
    relative liver weights were increased in the groups exposed during
    days 6-15 and 8-15. After exposure during days 6-15, ALAT levels
    were significantly increased in pregnant and non-pregnant animals,
    the pregnant animals showing the smaller increase. Among the
    controls, no difference in ALAT activity was observed. An increase

    in total litter resorptions was observed after exposure through days
    8-15. Mean fetal body weight and crown-rump length were decreased
    significantly if the dams had been exposed through days 1-7 or 6-15
    of pregnancy. In the exposed groups an increased number of fetuses
    with delayed ossification of skull bones and sternebrae was
    observed, especially in the days 1-7 and 8-15 exposed groups. The
    incidence of cleft palates significantly increased in fetuses from
    dams exposed to 490 mg/m3 through days 8-15 of gestation for 7 h
    each day.

         Published information for embryotoxicity and teratogenicity of
    chloroform in rat, mouse and rabbit by oral and inhalation exposure
    are summarized in Table 11.

    7.6  Mutagenicity and related end-points

         Very many genotoxicity assays have been conducted with
    chloroform and the data currently available are summarized in Tables
    12 and 13. Some of these reports are from a large collaborative
    study comparing intra-laboratory variations in testing methodology
    (De Serres & Ashby, 1981).

         Two problems potentially compromise the interpretation of
    mutagenicity data on chloroform. First, there is a possibility that
    ethyl and diethylcarbonate, produced by reaction of phosgene with
    ethanol that is routinely added to U.S.P (US Pharmacopoeia)
    chloroform, could generate false positive results. Secondly, testing
    of chloroform must be done in a sealed system because of its
    volatility, and so studies that did not take this factor into
    account could give false negative results.

         In data presented in Table 12, three separate studies using the
    Ames assay were conducted under sealed conditions to assure
    chloroform retention. All three studies yielded negative test
    results.

         In not all studies was it reported whether an  in vitro assay
    was performed in a sealed chamber to prevent chloroform evaporation
    (Table 12). However, dimethylsulfoxide (DMSO) was often used as a
    solvent, thus increasing retention in the media. Furthermore, even
    in the case of an unsealed chamber, chloroform would be expected to
    stay in the media for a period of hours, and very high doses (up to
    10 mg/plate) were often used.

         Chloroform has been tested by a number of authors in validated
    bacterial systems with  Salmonella typhimurium and  Escherichia
     coli and showed to be negative both with and without metabolic
    activation. Only in one uncommon test with  Photobacterium
     phosphorum was a positive effect found (Wecher & Scher, 1982).


        Table 11.  Embryotoxicity, fetotoxicity and teratogenicity produced in animals by exposure to chloroform
                                                                                                                            
    Species                Dose                 Gestational days    Route of              Result            Reference
                                                  administered      administration
                                                                                                                            

    Rat           30, 100, 300 ppm                    6-15          inhalation          embryotoxic    Schwetz et al. (1974)
                                                    (7 h/day)                           fetotoxic

    Rat           20, 50, 126 mg/kg per day           6-15          oral                fetotoxic      Thompson et al. (1974)

    Rat           100, 200, 400 mg/kg per day         6-15          oral                fetotoxic      Ruddick et al. (1983)

    Mouse         100 ppm                             1-7,          inhalation          embryotoxic    Murray et al. (1979)
                                                      6-15,                             fetotoxic
                                                      8-15
                                                    (7 h/day)

    Rabbit        20, 35, 50 mg/kg per day            6-18          oral                fetotoxic      Thompson et al. (1974)

                                                                                                                            

    Table 12.  Mutagenicity studies with chloroform
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Bacterial systems

    Salmonella            TA1535        base-pair substitution    14C-labelled compound          + ra          i.m.,     -     Uehleke et al.
     typhimurium          TA1538        frame-shift mutation      tested; no further                          i.n.r.           (1976, 1977)
                                                                  details reported

    S. typhimurium        TA1535        base-pair substitution    5 mM tested; incubation         + m           PB       -     Uehleke et al.
                          TA1538        frame-shift mutation      in closed containers                                         (1976, 1977)
                                                                  (survival > 80%)

    S. typhimurium        TA98          frame-shift mutation      suspension test and              -          i.n.r.     -     Simmon et al.
                          TA1537        frame-shift mutation      vapour test; concentration      + r                    -     (1977)
                          TA1538        frame-shift mutation      in suspension test was 10%
                          TA100         base-pair substitution    v/v, no further details
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      up to 3600 µg/plate;             -                     -     Gocke et al.
                          TA1537        frame-shift mutation      incubation in air-tight         + r           PCB      -     (1981)
                          TA1538        frame-shift mutation      desiccators
                          TA100         base-pair substitution
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      test conditions not              -                     -     Trueman (1981)
                          TA1537        frame-shift mutation      reported                        + r           PCB      -
                          TA1538        frame-shift mutation
                          TA100         base-pair substitution
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      test conditions not              -                     -     Ichinotsubo
                          TA100         base-pair substitution    reported                         +                     -     et al. (1981b)

                                                                                                                                              

    Table 12 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    S. typhimurium        TA98          frame-shift mutation      0.5, 1.0, 5.0, 100,              -                     -     Venitt & Crofton-
                          TA100         base-pair substitution    200, 500 µg/plate               + r           PCB      -     Sleigh (1981)

    S. typhimurium        TA98          frame-shift mutation      microtitre fluctuation           -                     -     Gatehouse
                          TA1537        frame-shift mutation      test; 1, 5, 10 µg/ml            + r           PCB      -     (1981)
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      fluctuation test; 1-500          -                     ±     Hubbard et
                          TA100         base-pair substitution    µg/ml (not specified)           + r                    -     al. (1981)

    S. typhimurium        TA98          frame-shift mutation      solvent DMSO; no further         -                     -     Baker &
                          TA1537        frame-shift mutation      details                         + r           PCB      -     Bonin (1981)
                          TA1538        frame-shift mutation
                          TA100         base-pair substitution
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      test conditions not              -                     ±     Garner et al.
                          TA1537        frame-shift mutation      reported                        + r           PB       -     (1981)
                          TA100         base-pair substitution
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      50, 100, 200, 1000,              -                     -     MacDonald
                          TA1537        frame-shift mutation      2000, 5000 µg/plate             + r           PCB      -     (1981)
                          TA100         base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      solvents DMSO; no further        -                     -     Nagao &
                          TA1537        frame-shift mutation      details                         + r           PCB      -     Takahashi
                          TA100         base-pair substitution                                                                 (1981)

                                                                                                                                              

    Table 12 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    S. typhimurium        TA98          frame-shift mutation      0.1, 1.0, 10, 100, 500,          -                     -     Rowland &
                          TA1537        frame-shift mutation      2000 µg/plate; solvent          + r           PCB      -     Severn (1981)
                          TA1538        frame-shift mutation      DMSO
                          TA100         base-pair substitution
                          TA1535        base-pair substitution

    S. typhimurium        TA1535        base-pair substitution    10, 100, 1000, 10 000            -                     -     Richold &
                          TA1537        frame-shift mutation      µg/plate; solvent               + r           PCB      -     Jones (1981)
                          TA1538        frame-shift mutation      DMSO

    S. typhimurium        TA98          frame-shift mutation      test conditions not              -                     -     Simmon &
                          TA1537        frame-shift mutation      reported                        + r           PCB      -     Shepherd
                          TA1538        frame-shift mutation                                                                   (1981)
                          TA100         base-pair substitution
                          TA1535        base-pair substitution

    S. typhimurium        TA98          frame-shift mutation      solvent DMSO or water;           -                     -     Brooks &
                          TA1537        frame-shift mutation      0.2, 2, 20, 200, 2000           + r           PCB      -     Dean (1981)
                          TA1538        frame-shift mutation      µg/plate
                          TA100         base-pair substitution
                          TA1535        base-pair substitution
                          TA92          interstrand DNA 
                                        crosslinks

    S. typhimurium        TA98          frame-shift mutation      10, 100, 1000, 10 000            -                     -     Van Abbé et
                          TA1537        frame-shift mutation      µg/plate                        + r           PCB      -     al. (1982)
                          TA1538        frame-shift mutation                                      + m           PCB      -
                          TA100         base-pair substitution

    S. typhimurium        TA1535        base-pair substitution    vapour test; exposure            -                     -     Van Abbé et
                          TA1538        frame-shift mutation      for 2, 4, 6 or 8 h              + r           PCB      -     al. (1982)

                                                                                                                                              

    Table 12 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    S. typhimurium        TM 677        forward mutation to       solvent DMSO; up to             + r           PCB      -     Skopek et al.
                                        azaguanine resistance     300 µg/ml                                                    (1981)

    Escherichia coli      WP2 uvrA      reversion to trp+         10, 100, 1000 µg/plate          + r           PCB      -     Gatehouse (1981)

    E. coli               WP2 uvrA      reversion to trp+         test conditions not             + r           PCB      -     Matsushima et
                                                                  reported                                                     al. (1981)

    E. coli               WP2p          reversion to trp+         solvent acetone; 0.1, 1,        + r                    -     Kirkland et
                          WP2 uvrA                                10, 100, 1000, 10 000            -            PCB      -     al. (1981)
                                                                  µg/plate

    E. coli               WP2p          reversion to trp+         0.5, 1.0, 5, 10, 50,             -                     -     Venitt &
                          WP2 uvr-p                               100, 200, 500 µg/plate          + r           PCB      -     Crofton-Sleigh
                                                                                                                               (1981)

    E. coli               K12           base-pair substitution    14C-labelled compound          + ra          i.m.      -     Greim et al.
                                        (not specified)           tested; no further details                                   (1977)
                                                                  reported

    Photobacterium        PPL-          reversion to normal       disc-diffusion assay; no         -                     +     Wecher &
     phosphoreum                        light emission            further details reported                                     Scher (1982)

    Non-mammalian eukaryotic systems

    Allium cepa                         chromosomal aberrations   solvent: DMSO; 0, 250, 500,                            +     Cortés et al.
                                                                  1000, 1500, 2500, 5000,                                      (1985)
                                                                  10 000 µg/ml

    Saccharomyces         D7            mitotic gene conversion   no details reported           + n.r.        i.n.r.     -     Zimmermann &
     cerevisiae                         at trp 5 locus                                                                         Scheel (1981)

                                                                                                                                              

    Table 12 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    S. cerevisiae         D7            mitotic gene conversion   21, 41, 54 mM; incubation        -                     +     Callen et al.
                                        at trp 5; mitotic         in screw-capped glass                                        (1980)
                                        recombination at ade 2,   tubes
                                        reversion at ilv 1 loci

    S. cerevisiae         D6            mitotic aneuploidy        agar added                      + r           PCB      -     Parry & Sharp
                                                                                                                               (1981)

    S. cerevisiae         D6            mitotic aneuploidy        direct incubation in            + r           PCB      ±     Parry & Sharp
                                                                  plastic bottles, 25, 50,                                     (1981)
                                                                  100 µg/ml; idem in glass        + r           PCB      -
                                                                  bottles

    S. cerevisiae         JD1           mitotic gene conversion   up to 1000 µg/ml;               + r           PCB      ±     Sharp & Parry
                                        at trp 5 locus and his    incubation in plastic                                        (1981a)
                                        5 polaron                 containers

    S. cerevisiae         JD1           idem as above             idem as above, only             + r           PCB      -     Sharp & Parry
                                                                  incubation in glass                                          (1981a)
                                                                  containers

    S. cerevisiae         D4            mitotic gene conversion   0.33, 1.0, 3.33, 100,            -                     -     Jagannath et
                                        at ade 2 and trp 5 loci   333.3 µg/plate                  + r           PCB      -     al. (1981)

    S. cerevisiae         T1            mitotic crossing over     100, 1000 µg/plate               -                     -     Kassinova et
                          T2            at ade 2                                                  + r           PCB      -     al. (1981)

    S. cerevisiae         XV 185-14     reversion at his 1,       solvent DMSO; 111,               -                     -     Mehta & Von
                          C (haploid)   hom 3, and arg 4 loci     1111 µg/ml                    + n.r.        i.n.r.     -     Borstel (1981)

    Schizosaccharomyces   P1            forward mutation at ade   5, 7.5, 10 µg/ml                 -                     -     Loprieno
     myces pombe                        1, 3, 4, 5 and 9 loci                                     + r           PCB     (+)    (1981)

                                                                                                                                              

    Table 12 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Aspergillus           35            forward mutation          0.5% (survival 26%)              ?                     -     Gualandi
     nidulans             (haploid)     (induction of methionine                                                               (1984)
                                        suppressors)

    A. nidulans           P1            somatic segregation       0.5% (survival 16.5%)            ?                     -     Gualandi
                                        (crossing-over and non-                                                                (1984)
                                        disjunction)

    Drosophila            Berlin K      sex-linked recessive      Basc-test; 24 mM; adult                                -     Gocke et al.
     melanogaster         wild and      lethal test               feeding                                                      (1981)
                          Basc

    D. melanogaster       Berlin K      sex-linked recessive      Basc-test; solvent DMSO;                               -     Vogel et al.
                          wild and      lethal test               0.1, 0.2% treated at                                         (1981)
                          y mei 9a      25 °C for 3 days with
                          mei-41 D5     standard feeding 
                                        technique

    In vitro mammalian systems

    Chinese hamster       V79           forward mutation to       1, 1.5, 2, 2.5%                  -                     -     Sturrock
                                        8-azaguanine resistance                                                                (1977)

    Human                 lymphocytes   chromosome breakage       solvent acetone; 50, 100,       + r           PCB      -     Kirkland et
                                                                  200, 400 µg/ml                                               al. (1981)

    In vivo mammalian systems

    Mouse                 CD1           micronuclei in            intraperitoneal, 0.015,                                -     Tsuchimoto &
                                        polychromatic             0.03, 0.06 ml/kg body                                        Matter (1981)
                                        erythrocytes of bone      weight at 0 and 24 h
                                        marrow

                                                                                                                                              

    Table 12 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Mouse                 NMRI          micronuclei in            intraperitoneal, 238, 476,                             -     Gocke et al.
                                        polychromatic             952 mg/kg body weight at                                     (1981)
                                        erythrocytes of bone      0 and 24 h
                                        marrow

    Mouse                 B6C3F1        micronuclei in            intraperitoneal, about                                 -     Salamone et
                                        polychromatic             0.088 ml/kg body weight                                      al. (1981)
                                        erythrocytes of bone      at 0 and 24 h or at
                                        marrow                    0 h only

    Mouse                 ?             micronuclei in            route not reported; 100,                              (+)    Agustin &
                                        polychromatic             200, 400, 600, 700, 800,                                     Lim-Sylianco
                                        erythrocytes of bone      900 mg/kg body weight                                        (1978)
                                        marrow

    Rat                   Long-Evans    chromosomal aberrations   intraperitoneal,                                       +     Fujie et al.
                                        in bone marrow            1.2-119.4 mg/kg body                                         (1990)
                                        6-597 mg/kg body weight   weight; oral,

    Host-mediated assays

    S. typhimurium        TA1535        base-pair substitution    test conditions not                                    -     Agustin & Lim-
                                                                  reported                                                     Sylianco (1978)

    S. typhimurium        TA1537        frame-shift mutation      test conditions not                                    +     Agustin & Lim-
                                                                  reported                                                     Sylianco (1978)

                                                                                                                                              

     Table 12 (contd)

    a +    = with metabolic activation     b PB     = phenobarbital                  c +   = positive
      -    = without metabolic activation    PCB    = polychlorinated biphenyls        (+) = weakly positive
      m    = mouse                           i.m.   = intact microsomes added          ±   = equivocal; study cannot be evaluated
      r    = rat                             i.n.r. = inducer not reported             -   = negative
      ra   = rabbit
      n.r. = species not reported
      ?    = not reported if metabolic
             action was used

    Table 13.  Indicator studies with chloroform
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Bacterial systems

    Escherichia coli      WP2, WP67     DNA damage (growth        test concentrations not          -                     -     Tweats (1981)
                          uvrA pol      inhibition)               specified; no further           + r           PCB      -
                          A & CM 871                              details
                          uvrA lexA
                          recA

    E. coli               WP2, WP67     DNA damage                test conditions not              -                     -     Green (1981)
                          uvrA pol      reported
                          A & CM 871
                          uvrA lexA
                          recA

    E. coli               W3110         DNA damage                liquid suspension test;          -                     -     Rosenkranz et
                          (polA+),                                25 mg/ml; solvent DMSO          + r           PCB     (+)    al. (1981)
                          P3478                                   or water
                          (POLA1-)

    E. coli               JC 2921 rec   DNA damage                test conditions not              -             +       -     Ichinotsubo
                          JC 9238 rec                             reported                         +                     +     et al.
                          JC 8471 rec                                                                                          (1981a)
                          JC 5519 rec
                          JC 7689 rec
                          JC 7623 rec

    E. coli               56-161        induction of prophage     0.5, 5 mg/ml                    + r           PCB      -     Thomson (1981)
                          env A         lambda in lysogenic                                                              -
                          C 600         E. coli

                                                                                                                                              

    Table 13 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Bacillus subtilis     H17 rec+      DNA damage not further    maximum concentration:          + r                    -     Kada (1981)
                          M45 rec-      specified                 20 µl/plate                    + yf                    -
                                                                                                 + jc                    -

    Non-mammalian eukaryotic systems

    Allium cepa                         SCE                       solvent DMSO; 0, 250, 500,                             ±     Cortés et al.
                                                                  1000, 1500, 2500, 5000,                                      (1985)
                                                                  10 000 µg/ml

    Saccharomyces         T4            DNA repair                solvent DMSO; 0.1, 1.0%          -                     -     Kassinova et
     cerevisiae           T5                                                                                                   al. (1981)

    S. cerevisiae         197/2d        DNA repair                100, 300, 600, 750 µg/ml;        -                     ±     Sharp & Parry
                          rad 3,                                  incubation in plastic            +                     ±     (1981b)
                          rad 18,                                 bottles
                          rad 52,
                          trp 2

    In vitro mammalian systems

    Chinese hamster       ovary         SCE                       0.7% (after exposure, 78%       + r           PCB      -     White et al.
                                                                  of dose remained)                                            (1979)

    Chinese hamster       ovary         SCE                       0.01, 0.1 µg/ml; solvent        + r           PCB      -     Perry & Thomson
                                                                  DMSO                                                         (1981)

    Chinese hamster       ovary         SCE                       0.001, 0.01, 0.1 mM              -                    + d    Athanasiou &
                                                                                                                               Kyrtopoulos
                                                                                                                               (1981)

                                                                                                                                              

    Table 13 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Rat                   erythroblast  SCE                       only 1.0 mM tested               +                    (+)    Fujie et al.
                                                                                                                               (1993)

    Syrian hamster        embryo        adenovirus                0.12, 0.25, 0.50, 1.0,           -                     +     Hatch et al.
                                        transformation            2.0 ml/sealed chamber                                        (1983)
                                                                  (4.6 litre)

    Baby hamster          kidney        cell transformation       test conditions not              -                    (+)    Daniel & Dehnel
                                                                  reported                        + r                    -     (1981)

    Baby hamster          kidney        cell transformation       0.25, 2.5, 25, 250 µl/ml         -                     -     Styles (1979,
                                                                                                                               1981)

    Rat                   primary       UDS                       0.00084-8.4 mM                   -                     -     Althaus et al.
                          hepatocytes                                                                                          (1982)

    Mouse (B6C3F1)        primary       UDS                       0.01-10 mM                       -                     -     Larson et al.
                          hepatocyte                                                                                           (1994c)

    Human                 primary       UDS                       4 cases 0.01-1.0 mM              -                     -     Butterworth
                          hepatocyte                                                                                           et al. (1989)

    Human                 lymphocytes   SCE                       0.016-50 mM                      -                     +     Morimoto &
                                                                                                                               Koizumi (1983)

    Human                 lymphocytes   SCE                       solvent acetone; 25, 50,        + r           PCB      -     Kirkland et
                                                                  75, 100, 200, 400 µg/ml                                      al. (1981)

    Human                 lymphocytes   UDS                       0.1, 1.0, 10 mM                  -                     -     Perocco et
                                                                                                  + r           PB       -     al. (1983)

                                                                                                                                              

    Table 13 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Human                 lymphocytes   UDS                       2.5, 5, 10 µg/ml                 -                     -     Perocco &
                                                                                                  + r           PB       -     Prodi (1981)

    Human                 Hela cells    UDS                       0.1-100 µg/ml; solvent           -                     -     Martin &
                                                                  DMSO                            +r            PB       -     McDermid (1981)

    In vivo mammalian systems

    Mouse                 JCR/SJ        SCE in bone marrow        oral: 25, 50, 100,                                     +     Morimoto &
                                        cells                     250 mg/kg body weight                                        Koizumi (1983)
                                                                  per day for 5 days

    Mouse                 B6C3F1        DNA repair in liver       oral:  240 mg/kg body                                  -     Reitz et al.
                                                                  weight                                                       (1982)

    Mouse                 C57BL         sperm abnormalities       vapour exposure: 0.04,                                 +     Land et al.
                          x C3H                                   0.08%; 4 h/day for 5 days                                    (1981)

    Mouse                 CBA x         sperm abnormalities       intraperitoneal: 0.025,                                -     Topham (1980,
                          BALB/C                                  0.05, 0.075, 0.1, 0.25 mg/kg                                 1981)
                                                                  body weight per day for
                                                                  5 days; vehicle corn oil

    Rat                   F-344         UDS in hepatocytes        oral: 40, 400 mg/kg body                               -     Mirsalis et
                                                                  weight, single dose,                                         al. (1982)
                                                                  vehicle corn oil

    Mouse                 B6C3F1        UDS in hepatocytes        238, 477 mg/kg body weight,                            -     Larson et al.
                                                                  single dose, corn oil                                        (1994c)
                                                                  vehicle

                                                                                                                                              

    Table 13 (contd)
                                                                                                                                              

    Species               Strain/cells  Measured end-point        Test conditions              Activationa  Inducerb  Resultc  Reference
                                                                                                                                              

    Rat (neonatal)        liver and     DNA damage (3H elution)   vehicle corn oil; 200-400                              -     Petzold &
                          kidney cells                            mg/kg body weight                                            Swenberg (1978)

                                                                                                                                              

    a + = with metabolic activation; - = without metabolic activation; r = rat; yf = Yellowtail fish; jc = Japanese clam
    b PB = phenobarbital; PCB = polychlorinated biphenyls
    c + = positive; (+) = weakly positive; ± = equivocal; study cannot be evaluated; - = negative;
    d In this test chromosome aberrations were also reported to occur;  no details on this finding were reported.
    

         The majority of studies with non-mammalian eukaryotic systems
    (yeasts and other fungi) were negative. Positive results were
    obtained with  Saccharomyces cerevisiae D7, but only at the highest
    concentration tested, at which there was a marked toxic effect
    (Callen et al., 1980). It should be noted that this strain of yeast
    contains an endogenous cytochrome P450-dependent monooxygenase
    system. In  Schizosaccharomyces pombe an indication for a mutagenic
    effect was observed (Loprieno, 1981). The inconsistent results with
     Saccharomyces cerevisiae D6 and JD 1 were probably due to
    inadequate test conditions (exposure in plastic rather than glass
    containers) and therefore it can be considered that chloroform was
    non-mutagenic in these tests (Parry & Sharp, 1981; Sharp & Parry,
    1981a). In two sex-linked recessive lethal tests with  Drosophila
     melanogaster, no mutagenic activity was observed (Gocke et al.,
    1981; Vogel et al., 1981).

         Chloroform did not induce gene mutations in V79 Chinese hamster
    cells (Sturrock, 1977), or chromosomal aberrations in human
    lymphocytes  in vitro (Kirkland et al., 1981).

          In vivo mammalian testing comprised four micronucleus tests
    in mice, three of which gave a negative result (Tsuchimoto & Matter,
    1981; Gocke et al.,1981; Salamone et al., 1981). The fourth
    micronucleus test was reported to have given a weakly positive
    result (Agustin & Lim-Sylianco, 1978). The same authors found a
    positive effect in the mouse host-mediated assay with  Salmonella
     typhimurium TA1537 but not with TA1535.

         Indicator studies showed that chloroform induces
    sister-chromatid exchange (SCE) in hamster and human cells  in vitro
    in the absence of metabolic activation, and in mice  in vivo
    (Athanasiou & Kyrtopoulos, 1981; Morimoto & Koizumi, 1983). Positive
    or weakly positive results were reported in two tests on DNA damage
    and DNA repair with  Escherichia coli and  Saccharomyces cerevisiae
    (Sharp & Parry, 1981b; Rosenkranz et al., 1981; Ichinotsubo et al.,
    1981a).

         The ability of chloroform to induce unscheduled DNA synthesis
    (UDS) was examined in the  in vitro and  in vivo  hepatocyte DNA
    repair assays for the most sensitive site for tumour formation, the
    female mouse liver (NCI, 1976a,b). In the  in vitro assay, primary
    hepatocyte cultures from female B6C3F1 mice were incubated with
    concentrations from 0.01 to 10 mM chloroform in the presence of
    3H-thymidine. UDS was assessed by quantitative autoradiography. No
    induction of DNA repair was observed at any concentration. In the
     in vivo assay, animals were treated by gavage with chloroform in
    corn oil (238 and 477 mg/kg body weight). Primary hepatocyte
    cultures were prepared 2 and 12 h later, incubated with
    3H-thymidine and assessed for induction of UDS. No DNA repair
    activity was seen at either dose or at either time point. These

    negative results in the target organ are consistent with the
    suggestion that neither chloroform nor its metabolites react
    directly with DNA  in vivo.

         The ability of chloroform to induce DNA repair was examined in
    freshly prepared primary cultures of human hepatocytes from
    discarded surgical material. No activity was seen in cultures from
    four different individuals at concentrations as high as 1 mM
    chloroform (Butterworth et al., 1989).

         Given the large number of sensitive assays to which chloroform
    has been submitted, it is noteworthy that the reported positive
    responses are so few. Furthermore, these few positive responses were
    randomly distributed amongst the various assays with no apparent
    pattern or clustering for any test system. Taken together, the
    weight of evidence indicates that neither chloroform not its
    metabolites would appear to interact directly with DNA or possess
    genotoxic activity. The conclusion is consistent with the lack of
    initiating activity of chloroform (see section 7.7.4).

    7.7  Carcinogenicity

    7.7.1  Mice

         In a National Cancer Institute carcinogenicity study, B6C3F1
    mice received USP grade chloroform stabilized with ethanol (0.5-1%)
    in corn oil 5 times a week by gavage (NCI, 1976a,b). Dosing was
    stopped after 78 weeks and the animals were sacrificed after 92
    weeks. There were 20 animals per sex in the control group and 50
    animals per sex in the dosed groups. The dose levels changed after
    18 weeks, resulting in time-weighted average dose levels of 138
    (low) and 277 (high) mg chloroform/kg body weight for male mice and
    238 (low dose) and 477 (high dose) mg chloroform/kg body weight for
    female mice. Administration of the highest dose of chloroform
    reduced survival in the female mice. Causes of death were related to
    the observed liver tumours, pulmonary inflammation and cardiac
    thrombosis. This latter lesion was not observed in either the
    control or the low-dose group. Dose-related increased frequencies of
    hepatocellular carcinomas were found, the incidences being 1/18,
    18/50 and 44/45 at 0, 138 and 277 mg chloroform/kg body weight in
    the males and 0/20, 36/45 and 39/41 at 0, 238 and 477 mg chloroform
    per kg body weight in the females, respectively. Mice presented
    clinical signs of illness, i.e. a reduced food intake and an untidy
    appearance, but clear information on non-neoplastic lesions was not
    provided. There is evidence that tumour formation may have been
    secondary to induced cytolethality and regenerative cell
    proliferation (see Larson et al., 1994a, and section 7.2.1.1).

         Jorgenson et al. (1985) exposed female B6C3F1 mice to
    chloroform in their drinking-water for a period of two years. The

    concentrations were 0, 200, 400, 900 and 1800 mg/litre, and the
    numbers of animals were 430, 430, 150, 50 and 50 per group,
    respectively. Time-weighted average daily doses were 0, 34, 65, 130
    and 263 mg/kg body weight. Additional matched controls (50 animals)
    received the same quantity of drinking-water (without chloroform) as
    was consumed by the animals in the highest dose groups. Initially,
    25% of the animals in the two highest dose groups died, but later on
    the death rate was more or less equal to that in the control group.
    No treatment-related effects on either liver or total tumour
    incidence were observed. Lack of tumour formation is consistent with
    the lack of induced liver necrosis or regenerative hepatocyte cell
    proliferation when chloroform is administered in the drinking-water
    (see Larson et al., 1994a and section 7.2.1.1).

         The difference between the results obtained in the NCI study
    (1976a,b) and the Jorgenson et al. (1985) study is probably related
    to the manner in which the compound was administered. When given in
    the drinking-water, only small amounts of chloroform reach the
    liver, corresponding to each sip taken. Apparently, these small
    doses and delivery rates can be metabolized, detoxified and
    eliminated without liver damage (Larson et al., 1994a). When similar
    daily amounts are given as a single bolus dose in corn oil, it is
    probable that the high rate of delivery to the liver results in the
    production of toxic metabolites that overwhelm detoxification
    mechanisms, resulting in cell death and regenerative cell
    proliferation (Larson et al., 1994a). The choice of vehicle may also
    contribute to the observed difference in toxicity (Bull et al.,
    1986) (see also section 7.2.1.1).

         Roe et al. (1979) administered daily chloroform (British
    Pharmacopoeia quality) in a toothpaste base (vehicle) to ICI mice
    (control group 104 animals per sex, dose groups 52 animals per sex)
    by gavage, 6 days a week for 80 weeks, followed by a 16-week
    observation period. The dose levels were 0 (controls), 17 and 60
    mg/kg body weight. Mice that died during the first 15 weeks of the
    experiment were replaced by animals from a reserve group (which were
    probably also dosed, although this was not specified). The control
    toothpaste did not contain eucalyptol and peppermint oil, whereas
    the toothpaste containing chloroform did contain these substances.
    Treatment with chloroform resulted in slightly increased survival,
    especially in the males. The most common cause of death was
    respiratory failure. A slightly increased incidence of fatty
    degeneration was observed among the chloroform-treated animals.
    Total tumour incidence was increased in the male mice (20/37 and
    21/38 at 17 and 60 mg/kg body weight, respectively, versus 20/72 in
    the controls). Renal tumours (3 hypernephromas and 5 cortical
    adenomas) were reported in 8 out of 38 males of the high-dose group.

         In a second experiment by Roe et al. (1979), the influence of
    peppermint oil, eucalyptol and chloroform was determined separately.

    In this study, male ICI mice received 60 mg chloroform/kg body
    weight daily, in the same way as in the study reported above. The
    vehicle control (toothpaste without chloroform, eucalyptol and
    peppermint oil) and dose groups consisted of 260 and 52 male
    animals, respectively (the groups receiving a dose of peppermint or
    eucalyptol also consisted of 52 animals). Again, the survival in the
    chloroform-dosed group was better than in the control group. Total
    tumour incidence was lower in the chloroform-treated group (30/49
    versus 170/240 in the controls). However, administration of
    chloroform resulted in a kidney tumour frequency (hypernephromas and
    adenomas) of 9/49, compared with a control value of 6/240.

         In a third study by Roe et al. (1979), 60 mg chloroform/kg body
    weight in toothpaste (containing eucalyptol and peppermint oil) was
    administered daily to male mice (52 per group) of the ICI, CBA,
    C57BL and the CF1 strain for a period of 80 weeks. The chemical
    was also administered in arachis oil to male mice of the ICI strain.
    Each strain had its own control group. Terminal sacrifice was at 93,
    97-99, 104 and 104 weeks for the CF1, ICI, C57BL and CBA strains,
    respectively. In this study, a treatment-related increase in the
    survival was found in all strains tested, except for the CF1
    strain. Treatment with chloroform resulted in a higher incidence of
    renal changes in the CBA and CF1 strains but not in the C57BL
    strain. The cause of death in all four strains was renal neoplasia
    in combination with respiratory and renal disease. In the C57BL, CBA
    and CF1 strains no changes in tumour frequencies were observed. In
    the ICI mice, after treatment with chloroform in either the
    toothpaste vehicle or arachis oil, an increase in the incidence of
    malignant kidney tumours was found (3/47 versus 0/49 in the
    controls, toothpaste vehicle; 9/48 versus 0/50 in the controls,
    arachis oil vehicle).

         Though full results are not yet available, an additional
    carcinogenesis bioassay in which mice were exposed to chloroform by
    inhalation is under way (Matsushima, personal communication, 1993).

    7.7.2  Rats

         In a National Cancer Institute carcinogenicity study,
    Osborne-Mendel rats received USP grade chloroform stabilized with
    ethanol (0.5-1%) in corn oil 5 times a week by gavage (NCI,
    1976a,b). Dosing was stopped after 78 weeks and the animals were
    sacrificed after 111 weeks. There were 20 animals per sex in the
    control group and 50 animals per sex in the dosed groups. The dose
    levels changed after 23 weeks, resulting in time-weighted average
    dose levels of 90 (low dose) and 180 (high dose) mg chloroform/kg
    body weight for males and 100 (low) and 200 (high) mg chloroform/kg
    body weight for females. Administration of chloroform reduced
    survival in male and female rats in all dose groups. A clear
    pathological reason for this effect in the rats was not given. In

    male rats, dose-related increased frequencies of kidney epithelial
    tumours were observed (incidences: 0/19, 4/50 and 12/50 at 0, 90 and
    180 mg chloroform/kg body weight, respectively). In the females a
    non-significant increase in the frequency of thyroid tumours was
    found (incidences: 1/19, 8/49 and 10/46 at 0, 100 and 200 mg
    chloroform/kg body weight, respectively). Rats presented clinical
    signs of illness, i.e. a reduced food intake and an untidy
    appearance. However, clear information on non-neoplastic lesions was
    not provided.

         Reuber (1979) re-evaluated the histological sections of the NCI
    study (1976a,b) and reported the same neoplastic lesions as the NCI.
    In addition, he noted that chloroform-dosed female rats developed
    liver lesions that were not seen in the control females (i.e.
    cholangiofibromas 0/20, 1/39 and 3/39; cholangiocarcinomas 0/20,
    2/39 and 8/39; hyperplastic nodules 1/20, 7/39 and 12/39; and
    hepatocellular carcinomas 0/20, 2/39 and 2/39, for the control, low-
    and high-dose groups, respectively).

         Jorgenson et al. (1985) exposed male Osborne-Mendel rats via
    drinking-water to 0, 200, 400, 900 and 1800 mg chloroform/litre for
    a period of two years. Time-weighted average daily chloroform doses
    were 0, 19, 38, 81 and 160 mg/kg body weight and the numbers of
    animals were 330, 330, 150, 50 and 50 per group, respectively.
    Additional matched controls (50 animals) received the same quantity
    of drinking-water (without chloroform) as was consumed by the
    animals in the highest dose groups. As a probable consequence of
    reduced drinking and reduced body weights, death rate was reduced
    with increasing chloroform dosage and in the matched control group.
    The only dose-related effect was an increase in renal tubular cell
    adenomas and adenocarcinomas. The incidence for all kidney tumours
    was 5/301, 1/50, 6/313, 7/148, 3/48 and 7/50 for control, matched
    control and the 19, 38, 81 and 160 mg/kg groups, respectively. From
    38 mg/kg body weight upwards the increase in the frequency of all
    kidney tumours was statistically significant.

         In an inadequately reported study, Tumasonis et al. (1985)
    exposed male and female Wistar rats to 0 or 2900 mg chloroform per
    litre drinking-water during the lifetime of the animals. Animal
    numbers were 26 and 22 in the male and female control groups and 32
    and 45 in the male and female treated groups, respectively. The
    experiment started with weanlings. After 72 weeks, the
    drinking-water chloroform concentrations were reduced because of an
    increased intake of water by exposed animals. However, daily intakes
    of chloroform varied considerably and so the time-weighted average
    daily doses were estimated roughly from a figure in the report. They
    appeared to be around 180 mg/kg body weight in the males and around
    240 mg/kg body weight in the females. Body weights were decreased
    and life-span was increased in the exposed animals. A severe hepatic
    adenofibrosis (cholangiofibrosis) was observed in the exposed

    animals. Ten out of the 40 females examined showed hepatic
    hyperplastic nodules (none did in the control group). In the males
    no increase in the incidence of neoplastic nodules was found.

         Although full results are not yet available, an additional
    carcinogenesis bioassay in which rats were exposed to chloroform by
    inhalation is under way (Matsushima, personal communication, 1993).

    7.7.3  Dogs

         Heywood et al. (1979) administered chloroform to Beagle dogs at
    dose levels of 0, 15 and 30 mg/kg body weight (6 days/week) in
    toothpaste in a gelatin capsule for a period of 7.5 years. Sacrifice
    followed after an observation period of 19 to 23 weeks, during which
    the chloroform treatment was withdrawn. The control group consisted
    of 16 animals of each sex and the dose groups of 8 animals of each
    sex. There were no treatment-related increases in tumours.

    7.7.4  Studies on initiating-promoting activity

    7.7.4.1  Mice

         One week after a single intraperitoneal administration of
    ethylnitrosourea (0, 5 or 20 mg/kg body weight) to 15 days old CD1
    Swiss mice (both sexes), Pereira et al. (1985) exposed the animals
    to chloroform via drinking-water at concentrations of 0 or 1800
    mg/litre until they were 51 weeks of age, after which the animals
    were sacrificed. The chloroform treatment did not affect the liver
    or lung tumour frequency in the females and the lung tumour
    frequency in the males. However, the liver tumour frequency in the
    males appeared to be reduced after the treatment.

         Capel et al. (1979) administered chloroform as a drinking-water
    solution (estimated daily doses of 0, 0.15 or 15 mg/kg body weight)
    to male mice either from 14 days before or from 14 days before to 14
    days after intraperitoneal injection with Ehrlich ascites cells (TO
    strain), subcutaneous injection with B16 melanoma cells (C57BL
    strain) or intramuscular injection with Lewis lung carcinoma cells
    (C57BL strain). Chloroform treatment enhanced the growth of Ehrlich
    ascites cells (measured as intraperitoneal tumour cell DNA) at the
    high dose level. In comparison with the controls, more animals
    receiving chloroform at both dose levels had organs invaded with B16
    melanoma cells. Lewis lung tumour growth, measured as primary tumour
    size or pulmonary metastases, was not significantly enhanced at
    low-dose chloroform treatment, but after treatment with the high
    dose the number of pulmonary metastases and tumour size were
    markedly increased.

         In a two-stage (initiation/promotion) treatment protocol,
    Klaunig et al. (1986) studied the effect on liver tumour incidence

    in male B6C3F1 mice (35/group) after continuous treatment with 600
    and 1800 mg chloroform/litre drinking-water for 52 weeks to
    determine if chloroform expresses its hepatocarcinogenicity through
    tumour promotion mechanisms. Two groups received 600 and 1800 mg
    chloroform/litre drinking-water containing diethylnitrosamine (DENA;
    10 mg/litre) during the first 4 weeks of exposure. Two other groups
    received 600 and 1800 mg chloroform/litre drinking-water without
    DENA. The DENA groups constituted the initiated groups. One
    initiated and one non-initiated control group were included.
    Chloroform did not affect the incidence of liver or lung tumours by
    itself, and even inhibited liver and lung tumorigenesis in the
    DENA-initiated mice, compared with DENA treatment alone.

    7.7.4.2  Rats

         Deml & Oesterle (1985, 1987) studied the ability of chloroform
    to promote the development of liver tumours. Female Sprague-Dawley
    rats were initiated for liver tumours by administration of a single
    dose of 8 mg dimethyl nitrosamine/kg body weight. This was followed
    by administering chloroform (25, 100, 200 and 400 mg/kg body weight)
    in an olive oil vehicle twice weekly for 11 consecutive weeks. There
    was a dose-related increase of ATPase-negative, gamma-glutamyl
    transpeptidase (GGTase)-positive and glycogen-storing foci of cells
    within the liver. For example, ATPase-deficient foci were increased
    from approximately 2-fold to 5-fold by doses of 100 and 400 mg/kg,
    respectively. These data demonstrate that chloroform in an oil
    vehicle will probably promote development of hepatic tumours in
    rats.

         Herren-Freund & Pereira (1986) evaluated the ability of
    chloroform to act as an initiator, promoter and co-carcinogen in
    B6C3F1 mice and male Sprague-Dawley rats. In rats, the initiator
    was administered 18-24 h following a two-thirds partial hepatectomy.
    Diethylnitrosamine (0.5 mmol/kg body weight) was used as the
    positive control for initiation and phenobarbital (500 mg/litre
    drinking-water) was used as the positive control for promotion.
    Ethylnitrosourea (ENU) was the positive control for initiator in
    15-day-old mice and phenobarbital (500 mg/litre drinking-water) was
    used as the positive control for promotion. Chloroform was
    administered as a single dose of 180 and 360 mg/kg body weight as an
    initiator (no vehicle) in the rat and 1800 mg/litre drinking-water
    for 48 weeks as a promoter. There was no evidence that chloroform
    was able to act as an initiator in rats. Moreover, it did not act as
    a tumour promoter in either mice or rats, but actually decreased the
    numbers of hepatic tumours induced in neonatal mice by ENU.
    Concurrent administration of chloroform and DENA to the rat had no
    significant effect on foci or tumour development in rats. These data
    further suggest that the corn oil vehicle is important to the
    hepatocarcinogenic effects of chloroform.

         In a previous experiment, Pereira et al. (1982) had examined
    the effect of chloroform as an initiator and promoter. Chloroform
    was administered at 180 mg/kg body weight in a single dose as an
    initiator and 180 mg/kg body weight twice a week for 53 days as a
    promoter. In this case, tricaprylin was the vehicle. Chloroform had
    no activity as an initiator. There was a small, but statistically
    significant, increase in the numbers of GGTase-positive foci in the
    promotion study.

         Although chloroform is an established rodent carcinogen,
    several studies have shown that chloroform administered in impolar
    solvents also has anti-cancer properties as it inhibits tumour
    growth in mouse liver and in the gastrointestinal tract of the rat
    (Pereira et al., 1985; Daniel et al., 1989).

         Chloroform administered in drinking-water (0, 900 and 1800
    mg/litre) to Fischer-344 rats significantly decreased
    gastrointestinal (GI) tumours that were initiated by a single 200
    mg/kg dose of dimethyl hydrazine (DMH) (Daniel et al., 1989). GI
    tumour incidence was 14/39 in animals treated with DMH alone and
    5/39 and 5/40 in the groups in which DMH treatment was followed by
    900 and 1800 mg chloroform/litre, respectively, for 39 weeks.

         Chloroform also inhibits the propensity for three
    gastrointestinal tract carcinogens, benzo (a)pyrene (BAP),
    1,2-dimethylhydrazine (DMH) and methylnitrosourea (MNU), to induce
    nuclear anomalies in the proximal colon of B6C3F1 mice (Daniel et
    al., 1991). These authors found that in mice pre-adapted to 1800 mg
    chloroform/litre drinking-water for 30 days prior to the carcinogen
    administration the level of nuclear anomalies induced in the
    proximal colon was reduced by four-fold for BAP and two-fold for
    both MNU and DMH. In the duodenum, chloroform was effective at
    inhibiting unclear anomalies only for MNU.

         Reddy et al. (1992) demonstrated that chloroform inhibits the
    development of diethylnitrosamine-initiated, phenobarbital-promoted
    gamma-glutamyl transpeptidase and placental form
    glutathione- S-transferase-positive foci in the liver of male
    Fischer-344 rats. They suggested that chloroform exerts its focal
    inhibitory effect by selectively killing the putative initiated
    cells.

         The lack of initiating activity in these initiation-promotion
    assays supports the conclusion that chloroform is non-genotoxic
    (section 7.6), and also indicates that the carcinogenic action of
    chloroform is attributable to a non-genotoxic/cytotoxic mode of
    action (sections 7.2.1.1 and 7.7). Interestingly, more of the above
    studies reported that chloroform inhibited the growth or formation
    of precancerous or cancerous cells than those that reported that
    chloroform had promotional activity.

    7.8  In vitro studies

          In vitro studies frequently provide insight into how
    chemicals induce cytotoxic effects. However, at high concentrations
    (e.g., 5 mM and above), the solvent effects of chloroform on cell
    membranes complicate the interpretation of these experiments. The
    preparations that have been studied are precision-cut slices taken
    from the liver, primary hepatocytes suspensions and cultures.

         Azri-Meehan et al. (1992) studied the cytotoxic effects of
    chloroform in liver slices taken from phenobarbital-treated rats. No
    comparison was made with non-induced animals. Concentrations in the
    range of 0.5 to 1.6 mM induced loss of intracellular potassium and
    glutathione. Reduced mitochondrial function (measured as decreases
    in dye reduction) was observed in the same dose range. A
    concentration of 0.2 mM had no effect.

         Glende & Recknagel (1992) examined the ability of a number of
    chlorinated hydrocarbons to activate phospholipase A2, presumably
    through damage to calcium-binding sites in the endoplasmic
    reticulum. At doses that induce 30 to 70% release of cellular
    lactate dehydrogenase (i.e. 9.8 mM), chloroform did activate
    phospholipase A2. This concentration is similar to that necessary to
    destroy the calcium-binding capacity of the endoplasmic reticulum.

         O'Hara et al. (1991) examined the effects of chloroform on the
    viability of hepatocytes in suspension (measured by potassium
    retention). These hepatocytes were isolated from control
    phenobarbital-treated rats. The minimum concentration required to
    produce an effect on potassium retention decreased from 10 mM in
    control hepatocytes to 1 mM in hepatocytes obtained from induced
    animals.

         A number of studies of chloroform cytotoxicity in suspensions
    of rat hepatocytes have been reported (Stacey, 1987). However, the
    very high nominal concentrations of chloroform that were apparently
    necessary to produce significant effects (i.e. 30 and 60 mM) raise
    considerable questions as to their relevance to  in vivo hepatic
    toxicity.

         An innovative approach has been developed for incubating
    hepatocyte suspensions with the chemical of interest, followed by
    observation of the cytotoxic response after placing the treated
    cells into culture (Kedderis et al., 1993a). Such cytotoxicity was
    observed when hepatocyte suspensions derived from B6C3F1 mice were
    incubated with concentrations of chloroform between 1.3 and 3.8 mM.
    These concentrations were consistent with peak liver concentrations
    expected with the high doses of chloroform utilized in the
    assessment of chloroform carcinogenicity in mice (NCI, 1976a,b), as
    predicted by the Corley et al. (1990) pharmacokinetic model

    (Kedderis et al., 1993b). The cytotoxicity of chloroform was
    potentiated by pretreating the mice with acetone to induce
    cytochrome P450 2E1.

         Although there have been substantial advances in the study of
     in vitro chloroform toxicity, the applicability of the results
    that are available to date to estimate hazards in humans remains to
    be established.

    7.9  Factors modifying toxicity; toxicity of metabolites

         The  in vivo toxicity of chloroform is modified by a range of
    factors. The rate of its biotransformation is a significant
    determinant of its toxicity. Hence, factors that increase or
    decrease chloroform biotransformation may alter the intensity of
    chloroform-induced toxicity. The activities of the cytochrome P450
    isoforms that catalyse the biotransformation of chloroform differ
    among species and between sexes of experimental animals. Moreover
    the activities of the enzymes that metabolize chloroform may be
    increased or decreased by exposure to chemicals, and exposure to
    chloroform itself may alter chloroform metabolism.

         In addition to differences in the rates of chloroform
    bioactivation, treatments that alter susceptibility are also
    important determinants of chloroform-induced toxicity. Cellular
    glutathione concentrations are an important determinant of
    susceptibility, and perturbations of glutathione homeostasis may
    affect markedly the toxicity of chloroform. Finally, for some of the
    treatments that alter chloroform toxicity discussed in this section,
    the mechanistic basis of these interactions is not well understood.

         Brown et al. (1974a) reported that inhalation exposure of
    phenobarbital-treated male Sprague-Dawley rats to chloroform at
    doses of 2.45 or 4.9 g/m3 (500 or 1000 ppm) for 2 h produced
    marked centrilobular necrosis that was accompanied by decreased
    hepatic glutathione concentrations.  in vitro studies showed that
    glutathione reduced the covalent binding of [14C]-chloroform
    metabolites to microsomal protein.

         Docks & Krishna (1976) observed that administration of
    chloroform decreased hepatic glutathione concentrations in
    phenobarbital-treated rats (male, Sprague-Dawley), but not in
    control animals 1 to 2 h after administration, and caused liver
    necrosis. Administration of isopropanol or acetone, which increased
    the covalent binding of chloroform metabolites (Sipes et al., 1973),
    did not alter hepatic glutathione concentrations.

         Starvation and carbohydrate restriction increase the  in vivo
    metabolism of chloroform and its hepato- and nephrotoxicity in rats
    (Nakajima & Sato, 1979; McMartin et al., 1981; Nakajima et al.,

    1982). In contrast, protein deficiency does not alter chloroform
    toxicity (McLean, 1970).

         Several authors have demonstrated that administration of
    alcohols, including ethanol (Kutob & Plaa, 1962b; Sato et al., 1980,
    1981), or ketones increases chloroform metabolism and
    hepatotoxicity. An extension of these studies to include a range of
    alcohols showed that methanol, ethanol, isopropanol,  tert-butanol,
    pentanol, hexanol, octanol and decanol all potentiate
    chloroform-induced liver injury and lower the LD50 of chloroform
    in male Sprague-Dawley rats (Ray & Mehendale, 1990). Aliphatic
    ketones, including acetone, 2-butanone, 2-pentanone, 2-hexanone,
    2,5-hexanedione, 2-heptanone and methyl isobutyl ketone, also
    increase chloroform-induced hepatotoxicity (Hewitt et al., 1990;
    Vézina et al., 1990), but treatment with 2-hexanone does not
    increase chloroform-dependent lipid peroxidation either  in vivo or
     in vitro (Cowlen et al., 1984a,b). The potentiating effect of
    alcohols and ketones in chloroform-induced hepatotoxicity is
    attributed to an increase in the activity of the cytochromes P450
    that metabolize chloroform (Koop et al., 1982; Ryan et al., 1986;
    Brady et al., 1989; Vézina et al., 1990).

         Harris et al. (1982) evaluated, by the intraperitoneal route,
    the toxicity of chloroform (0.2 ml/kg body weight) and carbon
    tetrachloride (0.1 ml/kg body weight) given alone or together to
    male rats. At these doses, neither chloroform nor carbon
    tetrachloride produced toxicity, but increases in SGPT activity and
    hepatic triglyceride and calcium concentrations were seen when both
    compounds were given together. Ikatsu & Nakajima (1992) showed that
    a single inhalation exposure to 490 mg/m3 (100 ppm) chloroform for
    8 h resulted in mid-zonal hepatotoxicity. In ethanol-treated rats
    exposed to both chloroform (50 ppm) and carbon tetrachloride (10
    ppm), liver necrosis and elevated plasma GOT/GPT activities were
    observed. These findings indicate that the toxicity of chloroform is
    elevated in the presence of carbon tetrachloride. O'Hara et al.
    (1991) studied the effect of chloroform and carbon tetrachloride in
    rat hepatocytes and demonstrated that the combined toxicity of both
    compounds was greater than additive.

         The pesticide kepone (chlordecone), but not its non-ketonic
    analogue mirex, increases chloroform-induced hepato- and
    nephrotoxicity (Hewitt et al., 1979, 1982; Iijima et al., 1983). In
    Mongolian gerbils, which are susceptible to chloroform-induced
    toxicity (200 or 500 µl/kg body weight, intraperitoneal), treatment
    with phenobarbital or chlordecone decreased the hepatotoxicity of
    chloroform (Ebel et al., 1987); in contrast, rats given 50 to 500
    µl/kg body weight chloroform (intraperitoneally) showed little
    hepatotoxicity, but toxicity was increased after treatment with
    phenobarbital or chlordecone.

         The drinking-water contaminants dichloroacetic acid (DCA) and
    trichloroacetic acid (TCA) potentiate chloroform toxicity (Davis,
    1992). Male and female rats (Sprague-Dawley) were orally treated
    with 0.92 or 2.45 mmol/kg body weight DCA or TCA and were given 0.75
    mg/kg body weight chloroform (intraperitoneally) 3 h later.
    Increases in plasma ALAT activities were observed in female, but not
    in male rats 24 and 48 h after giving DCA; in contrast, plasma ALAT
    activities were increased 24, but not 48 h, after giving TCA in both
    male and female rats. DCA administration increased blood urea
    nitrogen concentrations in female rats, but produced little effect
    in male rats, whereas TCA administration produced an effect only in
    female rats 48 h after treatment. The mechanism of the effect of DCA
    and TCA was not elaborated.

         Monochloroacetic acid (MCA) given by gavage to male (188 mg/kg
    body weight) or female (94 mg/kg body weight) Sprague-Dawley rats
    one hour before giving chloroform (520 mg/kg body weight)
    intraperitoneally increased chloroform-induced hepatotoxicity in
    male rats, but had little effect in female rats (Davis & Berndt,
    1992). Treatment with MCA alone decreased glomerular filtration
    rates in female rats. The mechanism by which MCA potentiated
    chloroform toxicity was not elucidated.

         Temporal variations in chloroform-induced hepatotoxicity have
    been observed in rats (Lavigne et al., 1983). Male Sprague-Dawley
    rats were given chloroform (0.5 ml/kg body weight) intraperitoneally
    at 9:00, 13:00, 17:00, 21:00 or 03:00 h and were killed 4 h after
    treatment. Hepatotoxicity, as assessed by serum GPT, GOT and LDH
    activities, was minimal and maximal at 09:00 h and 21:00 h,
    respectively, whereas glucose-6-phosphatase activity was decreased
    at 03:00 h and 13:00 h. When rats were starved for 16 h before
    giving chloroform at 09:00 h, toxicity was increased substantially.

         Charbonneau et al. (1991) studied the effect of acetone
    treatment on the toxicity with a range of binary mixtures of
    haloalkanes in rats. An increased hepatotoxic response was observed
    with binary mixtures containing chloroform, carbon tetrachloride,
    1,1,2-trichloroethane or 1,1-dichloroethylene.

         In vitamin-A-deficient rats, serum ALAT and particularly ASAT
    activities were increased after intraperitoneal administration of
    chloroform, compared to control rats (Savoure et al., 1992).

    8.  EFFECTS ON HUMANS

    8.1  Acute non-lethal effects

         Chloroform is irritating to mucous membranes, producing
    gastroenteritis with persistent nausea and vomiting. Symptoms
    following ingestion of chloroform are similar to those following
    inhalation (van der Heijden et al., 1986).

         Cases of severe intoxication after suicidal attempts, with the
    same pattern of symptoms as after anaesthetical use, have been
    reported by Schröder (1965). There are considerable inter-individual
    differences in susceptibility. Some persons presented serious
    illness after an oral dose of 7.5 g of chloroform, whereas others
    survived a dose of 270 g chloroform. The mean lethal dose for an
    adult is estimated to be about 45 g (Winslow & Gerstner, 1978).

         Rao et al. (1993) successfully managed acute toxicity from
    chloroform in a 33-year-old white woman who attempted suicide by
    injecting 0.5 ml of chloroform, and then drank half a cup the next
    morning. Plasma chloroform levels, measured by headspace GC,
    declined rapidly. Sequential measurement of biomarkers in serum for
    liver cell necrosis, liver function and liver regeneration indicated
    the presence of initial liver damage followed by recovery. The
    authors suggested that, in addition to biomarkers for liver
    necrosis, serial determinations of markers for liver regeneration
    provide objective evidence for recovery from chloroform poisoning.

         It has been reported that chloroform can cause severe toxic
    effects in humans exposed to 9960 mg/m3 (2000 ppm) for 60 min,
    symptoms of illness at 2490 mg/m3 (500 ppm) and can cause
    discomfort at levels below 249 mg/m3 (50 ppm) (Verschueren, 1983).

         Most data on the controlled exposure of man to chloroform have
    resulted from its clinical use as an anaesthetic. This use of
    chloroform was described as early as 1847 (Simpson, 1847). Induction
    of anaesthesia may result from inhalation of chloroform vapours at a
    concentration of 24 to 73 g/m3 air. For maintenance of
    anaesthesia, concentrations in the range of 12 to 48 g/m3 are
    required. As with animals, chloroform anaesthesia may result in
    death in humans due to respiratory and cardiac arrhythmias and
    failure. Because of the relatively high frequency of "late
    chloroform poisoning" (liver toxicity), its use as anaesthetic has
    been abandoned.

         Other effects related to chloroform inhalation are: increase in
    the rate and depth of respiration during induction and light
    anaesthesia, minute volume decrease in deep anaesthesia,
    hypothermia, depletion of adrenal adrenaline content, hypotension,
    depression of gastrointestinal tract motility, respiratory acidosis,

    hyperglycaemia, ketosis, constriction of the spleen, increase in the
    number of leucocytes (especially polymorphonuclear cells), a
    decrease in clotting time and an increase in prothrombin time. The
    characteristics and severity of the effects depend on depth and
    duration of anaesthesia (Adriani, 1970).

         The cardiac effects might be secondary and due to hypoxia,
    caused by depression of respiratory activity. No studies have been
    found in which this problem has been investigated in man (e.g., by
    forced respiration), but Taylor et al. (1976) obtained indications
    that chloroform itself produces cardiovascular disturbances in
    rabbits (viz. disturbances in left ventricular functioning and an
    increase in peripheral resistance; see section 7.1.2.3) after
    exposure to 244 mg/m3 for 1 min.

         In man, as well as in animals, renal tubular necrosis and renal
    dysfunction (anuria, proteinuria, uraemia, increase in blood urea
    nitrogen) have been observed (Kluwe, 1981). Recovering from
    chloroform anaesthesia, some patients may show the symptoms of a
    delayed chloroform poisoning several days later. Prostration,
    protracted nausea, vomiting, jaundice and coma due to hepatic
    dysfunction are observed. The patient may die within 5 days after
    anaesthesia. At autopsy, degeneration and necrosis of liver tissue
    have been found (Goodman & Gilman, 1970). In general the symptoms
    appear to be similar to those observed in animals.

         According to Oettel (1936) and Winslow & Gerstner (1978),
    exposure to concentrated chloroform vapours causes a stinging
    sensation in the eye. Splashing of the liquid into the eye evokes
    burning, pain and redness of the conjunctival tissue. Occasional
    injury of the corneal epithelium will recover fully within a few
    days. Dermal contact with chloroform causes chemical dermatitis
    (symptoms: irritation, reddening, blistering and burns).

    8.2  Epidemiology

    8.2.1  Occupational exposure

         Challen et al. (1958) reported the effects of exposure of
    workers (mostly female) to chloroform vapour in a factory during
    manufacture of lozenges containing the chemical. Eight workers (four
    working full-time and four half-time) were exposed to chloroform
    concentrations of 375 to 1330 mg/m3, with a peak concentration of
    5680 mg chloroform/m3, for periods of 3 to 10 years. The symptoms
    reported were lassitude, thirst, gastrointestinal distress, frequent
    and scalding urination, lack of concentration, depression and
    irritability. The management stated that some of the employees had
    been noticed staggering about at work. Nine other workers (one
    full-time, eight half-time), who were exposed to chloroform
    concentrations of 110 to 350 mg/m3 for 10 to 24 months, suffered

    from the same complaints as stated above, but to a lesser degree.
    Several liver function tests did not reveal signs of liver toxicity,
    but these tests were not very sensitive.

         Bomski et al. (1967) investigated the occurrence of hepatitis
    in a chemical factory in relation to the occurrence of this disease
    in the city where the factory was located. The 68 workers were
    exposed to occupational chloroform concentrations of 10 to 1000
    mg/m3 for 1 to 4 years. In this group of employees, a higher
    frequency of hepatitis was found than in the city inhabitants.
    Seventeen workers showed hepatomegaly and in three of them hepatitis
    was observed. Ten workers showed splenomegaly, but the cause of the
    splenomegaly was not discussed.

         The finding of a high frequency of hepatitis among
    occupationally chloroform-exposed workers, as compared to that in
    the city inhabitants, is supported by a recent report on a
    16-year-old patient who attempted suicide by ingesting chloroform.
    This led to the development of toxic hepatitis (Hakim et al., 1992).

         In a study by Phoon et al. (1975), the air in the workroom of
    13 persons with jaundice originally diagnosed as having viral
    hepatitis was analysed for chloroform. The chloroform concentration
    in the workroom appeared to be more than 1950 mg/m3. The period of
    exposure was less than 6 months. Because no worker had a history of
    fever and there was no relation to past medical history, it was
    concluded that the original diagnosis must have been wrong and
    should have been toxic jaundice. Five of the people with jaundice
    and four other colleagues had blood chloroform levels in the range
    of 1 to 2.9 mg/litre.

         In another factory 18 cases of what seemed to be hepatitis B
    were reported (Phoon et al., 1983). Investigation of the
    occupational environment revealed a constant exposure to chloroform,
    with concentrations in the range of 80 to 160 mg/m3. The exposure
    period of these workers was less than 4 months and the conclusion
    was drawn that these were cases of toxic jaundice related to
    chloroform exposure, because no infection with the hepatitis B virus
    could be established.

         A historical mortality study was carried out by Linde & Mesnick
    (1979). They investigated the cause of death of white male
    anaesthesiologists, who were occupationally exposed to chloroform
    vapours (extent of exposure was unknown). The death certificates
    used were of persons, who were presumed to be exposed during the
    1880-1890 period and who died during the 1930-1946 period.
    Comparison of their death certificates with those of several control
    groups did not exclude a possible association between cancer and
    chloroform exposure.

    8.2.2  General exposure

         There have been numerous reports over the last 15 years which
    have evaluated the relationship between chlorinated water and the
    incidence of cancer. Chloroform is but one of many by-products
    produced by reaction of chlorine with naturally occurring material
    in source waters (Bull & Kapfler, 1991). Many of these studies noted
    increased risk of cancer which at least partially fulfilled criteria
    for causality (e.g., consistency, specificity and temporal
    relationships).

         IARC (1991) reviewed the available studies and concluded the
    strongest evidence of increased risk related to exposure to
    chlorinated surface water relative to unchlorinated ground water for
    the incidence of cancer of the urinary bladder. However, the
    weight-of-the-evidence evaluation by IARC concluded that there is
    inadequate evidence for the carcinogenicity of chlorinated
    drinking-water in humans.

         Morris et al. (1992) conducted a meta-analysis which attempted
    to integrate quantitatively the results of previously published
    studies in which individual exposures were evaluated (i.e. case
    control and cohort studies). The authors identified increased rates
    of bladder and colo-rectal cancer in individuals exposed to
    chlorinated surface water, which appeared to exhibit a dose-related
    trend. Although this study was confounded by substantial differences
    in exposure variables that occur in different water supplies, higher
    risk rates were estimated when the analysis was restricted to those
    studies which were judged to have the highest quality exposure
    assessments. Because of the confounding of these results by chlorine
    residual levels and a multiplicity of other chemicals which are
    animal carcinogens and mutagens, none of the drinking-water studies
    specifically implicate chloroform as a human carcinogen.

         Kramer et al. (1992) studied the association between exposure
    to trihalomethanes in the water supply and adverse reproductive
    outcomes in the state of Iowa (USA). Estimations of chloroform
    exposure were based on municipal water surveys. After adjustment for
    maternal age, parity, prenatal care, marital status, education and
    maternal smoking, an increased risk for intrauterine growth
    retardation (abnormally low birth weight) was associated with
    chloroform concentrations above 10 µg/litre. Limitations of the
    study involve the ascertainment and classification of exposures to
    trihalomethanes (such as fluctuation of levels and exposure at
    individual level) and the influence of potential confounding
    influences of unmeasured contaminants.

    8.3  Abuse and addiction

         Exposure to chloroform may result in euphoria and therefore
    people expose themselves to chloroform by drinking the liquid or
    sniffing the vapours (Storms, 1973). Addiction to chloroform and
    chloroform-containing cough syrups has been reported by Heilbrunn et
    al. (1945) and Conlon (1963). According to Heilbrunn et al. (1945),
    addicts tolerated very high daily doses and presented neurological
    symptoms and degenerative changes in the brains.

         After an intravenous injection of 7.5 g of chloroform, a
    patient showed signs of pulmonary malfunction and haemolysis. In
    this case, kidney or liver toxicity was not reported (Timms et al.,
    1975).

    9.  EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND FIELD

    9.1  Freshwater organisms

         The data on the toxicity of chloroform to several freshwater
    organisms are listed in Table 14.

         Due to the volatility of chloroform, caution must be exercised
    in interpreting the test results, particularly those in open static
    systems where no chemical analysis of the actual concentration was
    carried out.

    9.1.1  Short-term toxicity

         The chemical is of low toxicity to unicellular plants and other
    microorganisms (concentration range of initial population growth
    inhibition: 125 to > 3200 mg/litre). Chloroform is moderately toxic
    to  Daphnia magna (LC50 = 29 mg/litre).

         The LC50 values for several species of fish are in the range
    of to 191 mg/litre. However, initial toxicity may occur at lower
    levels: the no-observed-lethal concentrations (NOLCs) for  Salmo
     gairdneri and  Lepomis macrochirus appear to be 8 and 3 mg/litre,
    respectively. At lower concentrations (< 13 mg/litre),  Salmo
     gairdneri shows loss of equilibrium, slow operculum movement and
    narcosis (Anderson & Lusty, 1980).

         In  Gasterosteus aculeatus, chloroform produced anaesthesia
    which could be maintained for at least 90 min at concentrations of
    210 mg/litre. Exposure to concentrations higher than 300 mg/litre
    resulted in decreased oxygen consumption and death (Jones, 1947),
    whereas concentrations lower than 120 mg/litre excited the animals
    and gave rise to considerable higher oxygen uptake.

         Chloroform is considerably more toxic to the juvenile stages of
    several species of amphibians. In a continuous-flow system, Birge et
    al. (1980) tested the toxicity of chloroform to embryo-larval stages
    of several species of amphibians after exposure for 7-9 days (Table
    14).  Hyla crucifer appeared to be the most susceptible species. An
    effect was found on the hatching rate of the embryos, which declined
    from 97% at 8 µg/litre to 4% at 7340 µg/litre. In addition there was
    some evidence of teratic larvae. During the 4 days post-hatching the
    LC50 declined from 760 to 270 µg/litre. The other species tested
    were less affected and only  Rana pipiens showed a high
    teratogenicity frequency in the offspring (100% at 27 mg/litre at
    18% hatching rate).


        Table 14.  Chloroform toxicity to water organisms
                                                                                                                                              

    Organism         Temperature  Medium               Stat/  Analysisc  Exposure    Parameter               Concentration  Reference
                        (°C)                           flowa             duration                             (mg/litre)
                                                                                                                                              

    Short-term toxicity

    Bacteria
    Pseudomonas        25         acc.d Bringmann &    S      -          16 h        initial reduction of       125         Bringmann &
    putida                        Kühn (1977)                                        cell multiplication                    Kühn (1977)

    Pseudomonas        25         acc.d Bringmann      S      -          16 h        initial change of          125         Bringmann (1973)
    fluorescens                   (1973)                                             culture turbidity

    Algae
    Microcystis        27         acc.d Bringmann      S      -          192 h       initial reduction of       185         Bringmann (1975)
    aeruginosa                    (1975)                                             cell multiplication

    Scenedesmus        25         acc.d Bringmann &    S      -          192 h       initial reduction of       1100        Bringmann &
    quadricauda                   Kühn (1977)                                        cell multiplication                    Kühn (1977)

    Haematococcus      20         acc.d Tümpling       S      -          4 h         10% reduction of           440         Knie et al. (1983)
    pluvialis                     (1972)                                             oxygen production

    Protozoans
    Entosiphon         25         Bringmann (1978)     S      -          72 h        initial reduction of       > 6560      Bringmann (1978)
    sulcatum                                                                         cell multiplication

    Uronema            25         Bringmann &          S      -          20 h        initial reduction of       > 6560      Bringmann &
    parduczi                      Kühn (1980)                                        cell multiplication                    Kühn (1980)

    Chilomonas         20         Bringmann et al.     S      -          48 h        initial reduction of       > 3200      Bringmann et al.
    paramaecium                   (1980)                                             cell multiplication                    (1980)

                                                                                                                                              

    Table 14 (contd)
                                                                                                                                              

    Organism         Temperature  Medium               Stat/  Analysisc  Exposure    Parameter               Concentration  Reference
                        (°C)                           flowa             duration                             (mg/litre)
                                                                                                                                              

    Crustaceans
    Daphnia magna      22         reconstituted well   S      -          48 h        LC50                       29          LeBlanc (1980)
                                  water, pH 7,
                                  hardness 173 mg
                                  CaCO3/litre

    Daphnia magna      19.8-20.9  lake water, pH       S      -          48 h        LC50                       65.7        Gersich et al.
                                  8.0, hardness                                                                             (1986)
                                  157 mg CaCO3/litre

    Daphnia magna      23         distilled water      S      -          48 h        LC50                       78.9        Abernethy et al.
                                                                                                                            (1986)

    Fish
    Cyprinus carpio    26         filtered well water  S      A          until       LC50                       97          Mattice et al.
    (mixed gametes)                                                      hatching                                           (1981)
                                                                         (3-5 days)

    Pimephales         25         carbon filtered      S      -          96 h        LC50                       129         Mayes et al.
    promelas                      lake water, pH                                                                            (1983)
    (10-15 days)                  7.6-8.3, hardness
                                  125 mg CaCO3/litre
    (30-35 days)       22         idem                 S      -          96 h        LC50                       171
    (60-100 days)      22         idem                 S      -          96 h        LC50                       103

    Brachydanio        20         dechlorinated        CF     -          48 h        LC50                       100         Slooff (1979)
    rerio                         tap water, pH 8,
                                  hardness 10 d.H.

                                                                                                                                              

    Table 14 (contd)
                                                                                                                                              

    Organism         Temperature  Medium               Stat/  Analysisc  Exposure    Parameter               Concentration  Reference
                        (°C)                           flowa             duration                             (mg/litre)
                                                                                                                                              

    Salmo              20         dechlorinated        CF     -          48 h        initial reduction          20          Slooff (1979)
    gairdneri                     tap water, pH 8,                                   of respiration
                                  hardness 10 d.H.                                   frequency

    Leuciscus          20         acc.d Mann (1975)    S      -          48 h        LC50                       162-191     Juhnke &
    idus melanotus                                                                                                          Lüdemann (1978)

    Carassius          5          aerated tap water    S      -          1 h         EC50                       97-167      Cherkin &
    auratus                                                                          (anaesthesia)                          Catchpool (1964)
                       20         aerated tap water    S      -          1 h         EC50                       167
                                                                                     (anaesthesia)

    Salmo              19         aerated river water  CF     A          96 h        LC50                       18          Anderson & Lusty
    gairdneri                                                                        NOLC                       8           (1980)

    Leopomis           19         aerated river water  CF     A          96 h        LC50                       18          Anderson & Lusty
    macrochirus                                                                      NOLC                       3           (1980)

    Micropterus        19         aerated river water  CF     A          96 h        LC50                       51          Anderson & Lusty
    salmoides                                                                        NOLC                       39          (1980)

    Ictalurus          19         aerated river water  CF     A          96 h        LC50                       75          Anderson & Lusty
    punctatus                                                                        NOLC                       68          (1980)

    Amphibians
    Hyla crucifer      20.5       acc.d Birge et al.   CF     A          until 4     LC50                       0.3         Birge et al. (1980)
    (eggs; 2 to                   (1979); pH 7.6,                        days after
    6h post-                      hardness 107 mg                        hatching 
    spawning)                     CaCO3/litre                            or death 
                                                                         (7 days in
                                                                         total)

                                                                                                                                              

    Table 14 (contd)
                                                                                                                                              

    Organism         Temperature  Medium               Stat/  Analysisc  Exposure    Parameter               Concentration  Reference
                        (°C)                           flowa             duration                             (mg/litre)
                                                                                                                                              

    Amphibians (contd)
                       20.5       idem                 CF     A          idem        NOLC                       0.009

    Rana pipiens       20.5       idem                 CF     A          idem (9     LC50                       4.2         Birge et al. (1980)
    (eggs; 30 min                                                        days in
    after                                                                total) 
    fertilization)     20.5       idem                 CF     A          idem        NOLC                       0.16

    Rana palustris     21.5       acc.d Birge et al.   CF     A          idem (8     LC50                       20.6        Birge et al. (1980)
    (eggs; 2 to                   (1979); pH 7.6,                        days in
    6 h post-                     hardness 104 mg                        total)
    spawning)                     CaCO3/litre
                       21.5       idem                 CF     A          idem        NOLC                       0.33

    Bufo fowleri       21.5       idem                 CF     A          idem (7     LC50                       35.1        Birge et al. (1980)
    (eggs; 2 to                                                          days in 
    6 h post-                                                            total)
    spawning)          21.5       idem                 CF     A          idem        NOLC                       0.33

    Long-term toxicity

    Fish
    Poecilia           22         Alabaster & Abram    Sb     -          14 days     LC50                       102         Könemann (1981)
    reticulata                                                                                                              (1964)

    Salmo              13.5 ± 1   acc.d Birge et al.   CF     A          until 4     LC50                       2.0         Birge et al. (1979)
    gairdneri                     (1979); pH 7.3,                        days after
    (eggs; 20 min                 hardness 48 mg                         hatching 
    after                         CaCO3/litre                            or death 
    fertilization                                                        (27 days 
                                                                         totally)

                                                                                                                                              

    Table 14 (contd)
                                                                                                                                              

    Organism         Temperature  Medium               Stat/  Analysisc  Exposure    Parameter               Concentration  Reference
                        (°C)                           flowa             duration                             (mg/litre)
                                                                                                                                              

                       13.5 ± 1   idem                 CF     A          idem        NOLC                       0.004

                       13.5 ± 1   idem, hardness 210   CF     A          idem        LC50                       1.24 
                                  mg CaCO3/litre

                       13.5 ± 1   idem                 CF     A          idem        NOLC                       0.003

                                                                                                                                              

    a S = static, CF = continuous flow; 
    b static conditions but test water changed every 24 h
    c A = concentration of test compound analysed during assay; - = no data
    d acc. = according to the medium described in these references
    

    9.1.2  Long-term toxicity

         Birge et al. (1979) tested the toxicity of chloroform for
    embryo-larval stages of  Salmo gairdneri (Table 14) after 27 days.
    The chemical was especially toxic for the unhatched embryos (LC50
    is about 2 mg/litre), but did not cause death in the larvae at
    concentrations up to 10.6 mg/litre. The occurrence of teratic
    survivors in the hatched population increased from 3% at 56 µg/litre
    to 40% at 10 mg/litre.

    9.2  Marine organisms

         The acute toxicity of chloroform to  Artemia salina was tested
    by Robinson et al. (1965). The observed effect was anaesthesia and
    the EC50 value was 68 mg/litre after 10 h of exposure in
    artificial sea water in closed containers under static conditions.
    The 50% immobilization concentration (IC50) of chloroform for
     Artemia salina nauplii, subjected to salinity stress, was
    determined in a static study using artificial sea water by Foster &
    Tullis (1985). The toxicity test began 30 h after hatching had
    commenced and lasted for 24 h. The IC50 was 37 mg/litre.

         Stewart et al. (1979) tested the acute toxicity of chloroform
    to larvae of  Crassostrea virginica. Chemical analysis showed a
    rapid decline of chloroform concentrations in the sea-water medium.
    The estimated LC50 was 1 mg/litre.

         Pearson & McConnell (1975) tested the acute toxicity to
     Limanda limanda in a continuous-flow system containing natural sea
    water and obtained an LC50 of 28 mg/litre.

         Cowgill et al. (1989) determined the sensitivity of the marine
    diatom  Skeletonema costatum to chloroform after exposure for 5
    days under static conditions. The EC50 values calculated were 477
    mg/litre and 437 mg/litre based on total cell count and total cell
    volume, respectively. The NOEC was 216 mg/litre.

    10.  EVALUATION OF HUMAN HEALTH RISKS AND EFFECTS ON THE ENVIRONMENT

    10.1  Evaluation of human health risks

    10.1.1  Exposure

         Based on estimates of mean exposure from various media, the
    general population is exposed to chloroform principally in food
    (approximately 1 µg/kg body weight per day), drinking-water
    (approximately 0.5 µg/kg body weight per day) and indoor air (0.3 to
    1 µg/kg body weight per day). Estimated intake from outdoor air is
    considerably less (0.01 µg/kg body weight per day). The total
    estimated mean intake for the general population is approximately 2
    µg/kg body weight per day. Available data also indicate that water
    use in homes contributes considerably to levels of chloroform in
    indoor air and to total exposure. For some individuals living in
    dwellings supplied with tap water containing relatively high
    concentrations of chloroform, estimated total intakes are up to 10
    µg/kg body weight per day.

         Workers may be exposed to chloroform during, for example, the
    production of chloroform itself, the synthesis of substances derived
    from chloroform (for example, chlorodifluoromethane), and the use of
    chloroform as a solvent, and also as a consequence of its formation
    in paper bleaching and sewage treatment facilities. For example,
    based on a national survey conducted from 1981 to 1983, NIOSH
    estimated that approximately 96 000 workers in the USA are
    potentially exposed to chloroform.

    10.1.2  Health effects

         The most important effects of chloroform are those on the liver
    and kidney. These effects are associated with the metabolism of
    chloroform to the reactive intermediate, phosgene. There are
    substantial interspecies and sex differences in the rates at which
    chloroform is metabolized. Data also indicate that reductive
    metabolism differs among species.

         The most universally observed toxic effect of chloroform is
    damage to the liver. The severity of these effects per unit dose
    administered depends on the species, the vehicle and the method by
    which the chloroform is administered. The lowest dose at which liver
    damage has been observed is 15 mg/kg body weight per day,
    administered to beagle dogs in a toothpaste base over a period of
    7.5 years. Effects at lower doses were not examined. Somewhat higher
    doses are required to produce hepatotoxic effects in other species.
    Although duration of exposure varied in these studies,
    no-observed-adverse-effect levels (NOAELs) ranged between 15 and 125
    mg/kg body weight per day.

         Effects on the kidney have been observed in male mice of
    sensitive strains and in F-344 rats. Severe effects have been
    observed in a particularly sensitive strain of male mice at doses as
    low as 36 mg/kg body weight per day.

         Daily 6-h inhalation of chloroform for 7 days consecutively
    induced atrophy of Bowman's glands and new bone growth in the nasal
    turbinates of F-344 rats. The NOEL for these effects was 14.7
    mg/m3 (3 ppm). The significance of these effects is being further
    investigated in longer term studies.

         The weight of the available evidence indicates that chloroform
    has little, if any, capability to induce gene mutation, chromosomal
    damage and DNA repair. There is some evidence of low-level binding
    to DNA, however. Chloroform does not appear capable of inducing
    unscheduled DNA synthesis  in vivo.

         Chloroform induced hepatic tumours in mice when administered by
    gavage in corn oil. However, when similar doses were administered in
    drinking-water to mice, hepatic tumours were not induced.

         The carcinogenic effects of chloroform on the mouse liver
    appear to be closely related to cytotoxic and cell replicative
    effects. The effects on cell replication paralleled variations in
    carcinogenic responses to chloroform due to vehicle and method of
    administration. It is of interest, in this regard, that chloroform
    administered in drinking-water was incapable of promoting, but
    rather inhibited, the development of liver tumours in mice.
    Chloroform does not appear capable of initiating liver tumours or
    inducing unscheduled DNA synthesis in the mouse liver. It would
    appear, therefore, that cytotoxicity followed by cell replication
    with prolonged administration of chloroform is associated with the
    development of liver tumours in mice.

         Chloroform induced kidney tumours in rats when administered by
    gavage in corn oil. However, results for this species were similar
    when the chemical was administered in the drinking-water.

         Experiments in F-344 rats have indicated that chloroform could
    cause damage and increase cell replication in the kidney at doses
    similar to those that induce renal tumours in Osborne-Mendel rats.
    These effects are produced by both oral (one single gavage) and
    7-day inhalation exposure. While these results are suggestive of an
    association, it is difficult to associate with any certainty the
    carcinogenic response with the toxic and replicative effects.
    Indeed, toxicity studies are short term and involve a rat strain
    that is unusually sensitive to the nephrotoxic effects of
    chloroform. This strain is different from that in which tumours were
    observed.

         There are some limited data to suggest that chloroform is toxic
    to the fetus, but only at doses that are maternally toxic.

    10.1.3  Approaches to risk assessment

         The following guidance is provided as a potential basis for the
    derivation of exposure limits by relevant authorities. By allocation
    of the tolerable and risk-specific intakes presented below based,
    for example, on the proportion of total intakes originating from
    each environmental medium presented in chapter 5, limits for
    exposure in drinking-water, food and air could be developed by local
    authorities (WHO, in press). However, local authorities may also
    wish to take into account local variations in the proportions of
    exposure from various media or factors such as cost, ease and
    effectiveness of control in order to develop risk management
    strategies appropriate for local circumstances. However, the
    ultimate objective should be reduction of total exposure from all
    sources to levels below the tolerable maximum intake and
    risk-specific intakes presented below. Moderate to short-term
    excedence of limits based on the guidance presented below does not
    necessarily imply significant risk to health and relevant public
    health authorities should be contacted before taking remedial
    action.

         Moreover, disinfection is unquestionably the most important
    step in the treatment of water for public supply. The paramount
    importance of microbiological quality requires some flexibility in
    the derivation of limits for exposure to chloroform in
    drinking-water. Where local circumstances require that a choice must
    be made between meeting microbiological limits or limits for
    disinfection byproducts, the microbiological quality must always
    take precedence. Efficient disinfection must never be compromised.

    10.1.3.1  Non-neoplastic effects

         The Task Group concluded that the data available are sufficient
    to develop a tolerable intake for non-neoplastic effects of
    chloroform on the basis of effects in animal species.

         The lowest effect level in long-term studies in animal species
    is that reported by Heywood et al. (1979) where slight
    hepatotoxicity (increases in hepatic serum enzymes and fatty cysts)
    was observed in beagle dogs that ingested 15 mg/kg body weight per
    day in toothpaste for 7.5 years. Liver fat content was also
    increased in B6C3F1 mice that ingested 34 mg/kg body weight per
    day in drinking-water for 2 years (Jorgenson et al., 1985). On the
    basis of these data, a tolerable daily intake (TDI) can be derived
    as follows:

           15 mg/kg body
           weight per day
    TDI = ----------------- = 0.015 mg/kg body weight per day
               1000           (15 µg/kg body weight per day)

    where:

    *    15 mg/kg body weight per day is the lowest-identified-effect 
         level (slight hepatotoxicity in the study on beagle dogs by 
         Heywood et al., 1979);

    *    1000 is the uncertainty factor (x 10 for interspecies
         variation,  x 10 for intraspecies variation and x 10 for use of
         an effect  level rather than a no-effect level).

         This value is likely to be conservative. It should be noted
    that no effects have been observed in adequate studies on other
    species exposed to higher doses administered in other vehicles.

    10.1.3.2  Neoplastic effects

         The Task Group concluded that the carcinogenic effects of
    chloroform should also be considered in the development of limits of
    exposure.

    a)  Liver tumours in female B6C3F1 mice

         Based on the available mechanistic data, the approach
    considered most appropriate for provision of guidance based on mouse
    liver tumours is division of a no-effect level for cell
    proliferation by an uncertainty factor. The NOEL for cytolethality
    and cell proliferation in B6C3F1 mice was 10 mg/kg body weight per
    day following administration in corn oil for 3 weeks (Larson et al.,
    1994a).

         On the basis of these data, a tolerable daily intake is derived
    as follows:

            10 mg/kg body
            weight per day
    TDI = ------------------ = 0.01 mg/kg body weight per day
                1000           (10 µg/kg body weight per day)

    where:

    *    10 mg/kg body weight per day is the NOEL for cytolethality  and
         cell proliferation in B6C3F1 mice observed in the short-term
         study of Larson et al. (1994a);

    *    1000 is the uncertainty factor (x10 for interspecies variation, 
         x10 for intraspecies variation and x 10 for severity of effect 
         (i.e. carcinogenicity) and less-than-chronic study).

    b)  Kidney tumours in male Osborne-Mendel rats

         Since data on cell proliferation are not available for the
    strain in which tumours were observed (Osborne-Mendel rats) and
    identified information on cell proliferation and lethality are short
    term (one single gavage and a 7-day inhalation exposure in F-344
    rats), it was considered premature to deviate from the default model
    (i.e. linearized multistage) as a basis for estimation of lifetime
    cancer risk.

         Based on the induction of renal tumours (adenomas and
    adenocarcinomas) in male rats in the study by Jorgenson et al.
    (1985), the total daily intake considered to be associated with a
    10-5 excess lifetime risk, calculated on the basis of the Global
    82 version of the linearized multistage model, is 0.0082 mg/kg body
    weight per day (8.2 µg/kg body weight per day). A body surface area
    correction was not incorporated due to the fact that chloroform is
    an indirect-acting carcinogen and that the rate of metabolism is
    similar in rodents and man.

    10.2  Evaluation of effects in the environment

         Chloroform may be released into the environment during its
    production, storage, transport and use. Significant amounts of
    chloroform may also enter the environment as a consequence of its
    formation during some chlorination processes (e.g., chlorination of
    water, paper bleaching).

         Chloroform is expected to volatilize readily from surface water
    and the surface of soils. It is also expected to be highly mobile in
    soils and may reach ground water.

         Chloroform has a residence time of several months in the
    atmosphere and can therefore be transported over long distances from
    the point of emission. Degradation by reaction with hydroxyl
    radicals is likely to be the only significant mechanism for
    decomposition of chloroform in the atmosphere. A half-life of around
    60 days has been estimated for this process.

         Chloroform appears to be resistant to biodegradation under
    aerobic conditions but is degraded under certain anaerobic
    conditions.

         Chloroform is toxic to the embryo-larval stages of some
    amphibian and fish species. The lowest reported LC50 is 0.3
    mg/litre (4- or 7-day exposure) for the embryo-larval stages of

     Hyla crucifer. It is less toxic to fish and  Daphnia magna. The
    LC50 values for several species of fish are in the range of 18 to
    191 mg/litre. There is little difference in sensitivity between
    freshwater and marine fish. The lowest reported LC50 for  Daphnia
     magna is 29 mg/litre (48-h exposure). Chloroform is of low
    toxicity to algae and other microorganisms.

         Levels of chloroform in surface water are generally low and
    would not be expected to present a hazard to aquatic organisms.
    However, higher levels of chloroform in surface water resulting from
    industrial discharges or spills may be hazardous to the
    embryo-larval stages of some aquatic species.

    11.  FURTHER RESEARCH

         A number of further studies is considered to be necessary:

    *    A study of compensatory cell regeneration in the liver and 
         kidney of the Osborne-Mendel rat

    *    Determination of reactive metabolite formation  in situ

    *    Studies on the mechanism of the species-specific
         carcinogenicity of chloroform including a) the identification 
         of the intermediate/metabolite responsible for the
         carcinogenicity of chloroform and b) its mode of action

    *    An inhalation carcinogenicity bioassay

    *    Further validation of PBPK models for chloroform with
         interspecies variations, including humans and dogs

    *    Further studies concerning the progression of nasal lesions in 
         the rat

    *    Additional long-term toxicity tests in aquatic organisms

    *     in vitro cytotoxicity/metabolism studies with human tissues

    12.  PREVIOUS EVALUATION BY INTERNATIONAL BODIES

         The International Agency for Research on Cancer evaluated
    chloroform in 1978 (IARC, 1979) and re-evaluated it in 1987 (IARC,
    1987). The conclusions were that there is inadequate evidence for
    the carcinogenicity of chloroform in humans but sufficient evidence
    for its carcinogenicity in experimental animals. The overall
    evaluation was that chloroform is possibly carcinogenic to humans
    (Group 2B).

         Chlorinated drinking-water was evaluated in 1990 (IARC, 1991)
    and the overall evaluation was that chlorinated drinking-water is
    not classifiable as to its carcinogenicity to humans (Group 3).
    Studies with chlorinated drinking-water gave no evidence for
    carcinogenicity of chloroform in humans (Group 3) (IARC, 1991).

         A drinking-water guideline value of 200 µg/litre for an excess
    lifetime cancer risk of 10-5 has been recommended for chloroform
    by the World Health Organization (WHO, 1993).

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    RESUME

         Le chloroforme se présente sous la forme d'un liquide volatil,
    limpide et incolore, à l'odeur caractéristique et au goût âcre et
    douceâtre. Il peut être décomposé par voie photochimique, il n'est
    pas inflammable et il est soluble dans la plupart des solvants
    organiques. Toutefois sa solubilité dans l'eau est limitée. Lors de
    la décomposition chimique, il peut y avoir formation de phosgène et
    d'acide chlorhydrique.

         Le chloroforme s'emploie dans certaines formulations de
    pesticides, comme solvant et comme intermédiaire dans la fabrication
    de certains dérivés. Son utilisation comme anesthésique ou dans des
    spécialités pharmaceutiques est interdite dans un certain nombre de
    pays. La production de chloroforme à des fins commerciales a atteint
    440 000 tonnes en 1987. Du chloroforme se forme également en
    quantités appréciables lors de la chloration de l'eau et du
    blanchiment de la pâte à papier.

         L'analyse de l'air, de l'eau et d'échantillons biologiques pour
    la recherche et le dosage du chloroforme peut s'effectuer selon
    plusieurs méthodes. La majorité d'entre elles consiste en une
    injection directe sur colonne, une adsorption sur un adsorbant
    activé ou une condensation dans un piège froid; on procède ensuite à
    une désorption ou à une extraction par un solvant qui est ensuite
    chassé avant analyse finale par chromatographie en phase gazeuse.

         On pense que la majeure partie du chloroforme présent dans
    l'eau finit par passer dans l'air, en raison de la volatilité de ce
    composé. Le temps de séjour du chloroforme dans l'atmosphère est de
    plusieurs mois et il en est éliminé après transformation chimique.
    Il résiste à la biodégradation aérobie par les bactéries du sol et
    des nappes phréatiques qui se développent sur des substrats
    endogènes ou en présence d'un supplément d'acétate. Il peut y avoir
    biodégradation en anaérobiose. La bioconcentration est faible chez
    les poissons d'eau douce. La dépuration est rapide.

         D'après l'estimation de l'exposition moyenne due aux divers
    milieux, on pense que la population générale est principalement
    exposée au chloroforme par l'intermédiaire de la nourriture, de
    l'eau de boisson et de l'air intérieur, dans des proportions à peu
    près égales. L'absorption estimative à partir de l'air intérieur est
    cependant beaucoup moindre. L'absorption moyenne totale estimative
    est d'environ 2 µg/kg de poids corporel, par jour. Les données
    disponibles indiquent également que l'utilisation domestique de
    l'eau contribue de façon très importante à la concentration du
    chloroforme dans l'air intérieur et par voie de conséquence à
    l'exposition totale. Pour certaines personnes qui vivent dans des
    habitations où l'eau de distribution renferme des concentrations 

    relativement élevées de chloroforme, on estime que l'absorption
    totale peut aller jusqu'à 10 µg/kg de poids corporel et par jour.

         Une fois administré par voie orale, le chloroforme est bien
    résorbé chez l'animal et l'homme, mais la cinétique d'absorption
    dépend du véhicule. Chez l'homme, après exposition par la voie
    respiratoire, 60 à 80% de la dose inhalée sont absorbés. Les
    principaux facteurs qui agissent sur la cinétique d'absorption du
    chloroforme après inhalation sont la concentration ainsi que la
    capacité de métabolisation, qui dépend de l'espèce. Chez l'homme et
    l'animal, le chloroforme est rapidement résorbé par la peau et l'on
    a montré qu'il pouvait être également absorbé par voie percutanée
    dans une proportion importante à partir de l'eau lors d'une douche.
    Il semble que l'hydratation de l'épiderme accélère la résorption du
    chloroforme.

         Le chloroforme se répartit dans l'ensemble de l'organisme.
    C'est dans les graisses, le sang, le foie, les reins, les poumons et
    le système nerveux que l'on trouve les plus fortes concentrations
    tissulaires. La répartition du chloroforme dépend de la voie
    d'exposition; la dose est plus forte dans les tissus
    extra-hépatiques après inhalation ou absorption percutanée qu'après
    ingestion. On a montré que chez plusieurs espèces animales et chez
    l'homme, le chloroforme pouvait traverser la barrière placentaire.
    Il s'élimine essentiellement dans l'air expiré sous forme de dioxyde
    de carbone. Non métabolisé, il demeure plus longtemps dans les
    graisses que dans les autres tissus.

         La biotransformation oxydative du chloroforme en
    trichlorométhanol est catalysée par le cytochrome P-450. Le
    trichlorométhanol produit, par élimination d'HCl, un intermédiaire
    réactif, le phosgène. Le phosgène peut être détoxifié en dioxyde de
    carbone par réaction avec l'eau ou en divers adduits par réaction
    avec des thiols, notamment le glutathion ou la cystéine. La réaction
    du phosgène sur les protéines tissulaires entraîne des lésions
    cellulaires et la mort. La liaison des métabolites du chloroforme à
    l'ADN est limitée. Le chloroforme peut également subir une
    biotransformation réductrice catalysée par le P-450, qui donne
    naissance au radical dichlorométhyl, lequel se fixe ensuite par
    liaison covalente aux lipides tissulaires. On n'a pas déterminé si
    cette biotransformation réductrice jouait également un rôle dans la
    cytotoxicité du chloroforme.

         Chez l'animal et l'homme exposés à du chloroforme, le
    chloroforme est éliminé d'une part sous forme de dioxyde de carbone
    et d'autre part sous forme inchangée. La fraction de la dose qui est
    éliminée sous forme de dioxyde de carbone varie avec cette dose et
    l'espèce en cause. La vitesse de biotransformation en dioxyde de
    carbone est plus élevée dans les microsomes hépatiques et rénaux des
    rongeurs (hamster, souris, rat) que dans ceux de l'homme. La

    biotransformation du chloroforme est également plus rapide dans les
    microsomes rénaux des souris que dans ceux des rats.

         En ce qui concerne la toxicité aiguë, c'est le foie qui est
    l'organe-cible chez le rat et plusieurs souches de souris. Les
    lésions hépatiques se caractérisent essentiellement, au début, par
    une infiltration graisseuse et une ballonisation des cellules, qui
    évoluent vers une nécrose centrilobulaire, puis une nécrose massive.
    Le rein est l'organe-cible chez les souris mâles appartenant à des
    souches plus sensibles. Au niveau du rein, les lésions débutent par
    une dégénérescence hydropigène qui évolue vers la nécrose des
    tubules proximaux. On n'a pas observé de toxicité rénale importante
    chez les femelles d'aucune souche de souris.

         La toxicité aiguë varie en fonction de la souche, du sexe et du
    véhicule. Chez la souris, la DL50 par voie orale varie de 36 à
    1366 mg/kg de poids corporel alors que chez le rat, elle peut aller
    de 450 à 2000 mg de chloroforme par kg de poids corporel. Après une
    seule exposition de 4 heures par voie respiratoire, on a observé des
    effets toxiques sur le foie chez la souris et le rat à des
    concentrations de chloroforme respectivement égales à 490 et 1410
    mg/m3.

         Ce sont les lésions du foie qui sont l'effet toxique du
    chloroforme le plus universellement observé. La gravité de ces
    effets par dose unitaire administrée dépend de l'espèce, du véhicule
    et du mode d'administration du chloroforme. La dose la plus faible à
    laquelle on ait observé ces lésions est de 15 mg/kg de poids
    corporel et par jour, administrée à des chiens "beagle" dans une
    base de pâte dentifrice, pendant une période de 7,5 années. On n'a
    pas recherché s'il y avait des effets à des doses plus faibles. Chez
    les autres espèces, les doses nécessaires pour produire des effets
    hépatotoxiques sont un peu plus élevées. Bien qu'au cours de ces
    différentes études, la durée d'exposition ait été variable, on a pu
    fixer la concentration sans effets nocifs observables à 15-125 mg/kg
    de poids corporel et par jour.

         Les effets au niveau du rein ont été observés chez des mâles
    appartenant à des souches sensibles de souris ainsi que chez des
    rats F-344. Ces effets étaient graves chez les mâles appartenant à
    une souche de souris particulièrement sensible, à des doses ne
    dépassant pas 36 mg/kg de poids corporel et par jour.

         Chez des rats F-344 à qui l'on avait fait inhaler du
    chloroforme 7 jours de suite, tous les jours pendant 6 heures, on a
    observé une atrophie des glandes de Bowman ainsi que la présence
    d'os néoformés dans les cornets du nez. La dose sans effets
    observables correspondante se situait à 14,7 mg/m3 (3 ppm). Des
    études à long terme se poursuivent afin d'évaluer la portée de ces
    effets.

         On a constaté l'apparition de tumeurs hépatiques chez des
    souris à qui l'on avait administré par gavage des doses quotidiennes
    de chloroforme dans de l'huile de maïs, à raison de 138 à 477 mg/kg
    de poids corporel. Toutefois, lorsque des doses analogues étaient
    administrées dans l'eau de boisson, le chloroforme était sans
    influence sur la proportion des tumeurs hépatiques qui se formaient
    chez ces souris. De plus, lors d'études sur le caractère promoteur
    éventuel de ce composé, on a observé, qu'administré dans l'eau de
    boisson, le chloroforme avait en fait une action inhibitrice sur la
    formation de tumeurs du foie provoquées chez la souris avec de la
    diéthylnitrosamine comme initiateur. Le véhicule utilisé ou la
    manière d'administrer le chloroforme conditionne donc de façon
    importante son pouvoir tumoro-inducteur au niveau du foie chez la
    souris.

         Le chloroforme a produit des tumeurs rénales chez des rats qui
    en avaient reçu quotidiennement par gavage, dans de l'huile de maïs,
    des doses allant de 90 à 200 mg/kg de poids corporel. Toutefois,
    chez cette espèce, les résultats se sont révélés analogues lorsque
    le produit était administré dans l'eau de boisson, ce qui indique
    que les effets ne dépendent pas entièrement du véhicule utilisé.

         Il semble que les effets cancérogènes du chloroforme sur le
    foie et le rein des rongeurs soient étroitement liés à son action
    cytotoxique ainsi qu'aux effets que ce composé exerce sur la
    réplication cellulaire dans les organes-cibles. On a constaté que
    ces derniers effets suivaient de près les modifications de la
    réponse cancérogène au chloroforme en fonction du type de véhicule
    et du mode d'administration. A la lumière des données disponibles,
    il semble que le chloroforme ne soit guère capable d'induire des
    mutations géniques ou d'autres types de lésions directes de l'ADN.
    En outre, le chloroforme ne semble pas non plus capable de jouer le
    rôle d'initiateur tumoral au niveau du foie chez la souris ni
    d'induire une synthèse non programmée de l'ADN  in vivo. En
    revanche, lorsqu'il est administré dans un véhicule huileux, le
    chloroforme peut se révéler un promoteur efficace des tumeurs
    hépatiques. Par conséquent, il est probable que, lors de
    l'administration prolongée de chloroforme, la cytotoxicité de ce
    composé et la prolifération cellulaire qu'il détermine sont les
    causes les plus importantes de la formation de tumeurs hépatiques et
    rénales chez les rongeurs.

         On dispose de quelques données limitées selon lesquelles le
    chloroforme serait toxique pour le foetus, mais uniquement à des
    doses auxquelles il est également toxique pour la mère.

         En général, le chloroforme détermine les mêmes symptômes
    toxiques chez l'homme que chez l'animal. Chez l'homme, l'anesthésie
    peut entraîner la mort par suite d'arythmie et d'insuffisance
    respiratoire et cardiaque. On a également observé chez l'homme une

    nécrose des tubules rénaux et une insuffisance rénale. Les doses les
    plus faibles pour lesquelles des cas de toxicité hépatique due à une
    exposition professionnelle au chloroforme ont fait l'objet de
    rapports, se situaient dans les limites de 80 à 160 mg/m3 (durée
    d'exposition de moins de 4 mois) selon une étude et allaient de 10 à
    1000 mg/m3 (durée d'exposition: 1 à 4 ans) selon une autre étude.
    On estime que la dose mortelle moyenne par voie orale pour un adulte
    est d'environ 45 g, mais on note d'importantes différences de
    sensibilité selon les individus. On est fondé à croire, selon
    certaines études épidémiologiques, qu'il existe une association
    entre l'exposition aux sous-produits des désinfectants présents dans
    l'eau de boisson et les cancers colorectaux ou vésicaux. Cependant,
    ces études souffrent de la présence de facteurs de confusion, entre
    autres faiblesses. Les preuves avancées à l'appui de la
    cancérogénicité pour l'homme de l'eau de boisson chlorée, sont
    insuffisantes. En outre, la présence de sous-produits des
    désinfectants utilisés ne peut être attribuée au chloroforme
    lui-même.

         Le chloroforme est toxique pour les stages embryo-larvaires de
    certaines espèces d'amphibiens et de poissons. La CL50 la plus
    faible dont il ait été fait état, se situait à 0,3 mg/litre pour les
    stades embryo-larvaires de  Hyla crucifer. Le chloroforme est moins
    toxique pour les poissons et pour la daphnie  Daphnia magna. Pour
    plusieurs espèces de poissons, les valeurs de la CL50 se situent
    dans les limites de 18 à 191 mg/litre. Il n'y a guère de différences
    de sensibilité entre les poissons d'eau douce et les poissons de
    mer. En ce qui concerne  Daphnia magna, la valeur la plus faible de
    la CL50 qui ait été signalée, était de 29 mg/litre. Le chloroforme
    est peu toxique pour les algues et autres microorganismes.

         Le Groupe de travail a estimé que les données disponibles
    étaient suffisantes pour établir une dose journalière tolérable
    (DJT) pour les effets non cancérogènes du chloroforme, ainsi qu'une
    dose spécifiquement liée au risque d'effets cancérogènes, sur la
    base des études effectuées chez l'animal; les valeurs ainsi fixées
    serviront de guide pour l'établissement de limites d'exposition par
    les autorités compétentes. Cependant, il est rappelé que lorsque les
    conditions locales imposent un choix entre le respect des limites
    microbiologiques ou celles qui concernent la présence de
    sous-produits de désinfection tels que le chloroforme, c'est la
    qualité microbiologique qui doit toujours l'emporter. Il ne faut
     jamais transiger sur l'efficacité de la désinfection.

         En se fondant sur l'étude de Heywood et al. (1979) et en
    introduisant un facteur d'incertitude de 1000 (x10 pour les
    variations interspécifiques, x10 pour les variations
    intraspécifiques et x10 pour l'utilisation d'une dose avec effet
    plutôt que d'une dose sans effet lors d'une étude subchronique), on
    obtient une DJT de 15 µg/kg de poids corporel; il faut rappeler que

    cette étude avait révélé l'existence d'une légère hépatotoxicité (à
    savoir une augmentation des enzymes hépatiques sériques et des
    kystes graisseux) chez des chiens "beagle" à qui l'on avait fait
    ingérer pendant 7,5 ans, une pâte dentifrice contenant du
    chloroforme à la dose de 15 mg/kg de poids corporel et par jour.

         En se fondant sur ce que l'on sait du mécanisme de ces
    phénomènes, la méthode que l'on juge la mieux adaptée pour obtenir
    une valeur-guide consiste à diviser la valeur de la concentration
    sans effet observable sur la prolifération cellulaire par un certain
    facteur d'incertitude. C'est ainsi que si l'on utilise la valeur de
    la dose sans effets observables obtenue par Larson et al. (1993b)
    pour la cytoléthalité et la prolifération cellulaire chez des souris
    B6C3F1 qui avaient reçu pendant 3 semaines, dans de l'huile de
    maïs, une dose quotidienne de chloroforme équivalant à 10 mg/kg de
    poids corporel, et en introduisant un facteur d'incertitude de 1000
    (x10 pour les variations interspécifiques, x10 pour les variations
    intraspécifiques et x10 pour la gravité de l'effet, c'est-à-dire la
    cancérogénicité et parce qu'il s'agit d'une étude subchronique), on
    obtient une DJT de 10 µg/kg de poids corporel.

         On admet que les tumeurs rénales observées chez le rat peuvent
    également être liées à l'action létale du chloroforme sur les
    cellules et à ses effets sur leur prolifération. Cependant, étant
    donné que l'on ne possède pas de données sur la prolifération
    cellulaire chez les souches où l'on a observé des tumeurs et qu'en
    outre, ce que l'on peut savoir de cet effet et de l'effet létal du
    chloroforme sur les cellules n'a été observé qu'à court terme (un
    seul gavage et une exposition par voie respiratoire de 7 jours), on
    estime qu'il est prématuré de s'écarter du modèle par défaut
    (c'est-à-dire multistade linéarisé) pour l'estimation du risque de
    cancer sur la durée de vie. D'après l'étude de Jorgenson et al.
    (1985) qui portait sur l'induction de tumeurs rénales (adénomes et
    adénocarcinomes), on a fixé à 8,2 µg/kg de poids corporel et par
    jour, la dose quotidienne totale jugée capable de produire un excès
    de risque de 10-5 sur toute la durée de la vie.

         La concentration de chloroforme dans les eaux de surface est
    généralement faible et ne semble pas présenter de danger pour les
    organismes aquatiques. Toutefois, la décharge ou le déversement de
    produits industriels pourrait entraîner la présence de
    concentrations plus élevées de chloroforme dans ces eaux et les
    rendre dangereuses pour les stades embryo-larvaires de certaines
    espèces aquatiques.

    RESUMEN

         El cloroformo es un líquido transparente, incoloro y volátil,
    con un olor característico y un sabor dulce ardiente. Se degrada
    fotoquímicamente, no es inflamable y es soluble en la mayor parte de
    los disolventes orgánicos. Sin embargo, su solubilidad en agua es
    limitada. Por degradación química del mismo pueden formarse fosgeno
    y ácido hidroclorhídrico.

         El cloroformo se utiliza en la formulación de plaguicidas, como
    disolvente y como intermedio químico. Su utilización como anestésico
    y en especialidades farmacéuticas está prohibida en algunos países.
    La producción comercial ascendió a 440 000 toneladas en 1987.
    También se producen cantidades apreciables de cloroformo en la
    cloración del agua y en el blanqueado de la pasta papelera.

         Existen varios métodos analíticos para determinar la presencia
    de cloroformo en el aire, el agua y los materiales biológicos. La
    mayor parte de esos métodos se basan en la inyección directa en
    columna, la adsorción en adsorbentes activados o la condensación en
    una cámara fría y posteriormente la desorción o evaporación mediante
    la extracción por disolventes o el calentamiento y el subsiguiente
    análisis por cromatografía de gases.

         Se supone que la mayor parte del cloroformo presente en el agua
    se transfiere finalmente al aire debido a su volatilidad. El
    cloroformo tiene un tiempo de residencia en la atmósfera de varios
    meses y desaparece de la misma por transformación química. Es
    resistente a la biodegradación por la población microbiana aeróbica
    de los suelos y de las capas acuíferas que viven en subestratos
    endógenos o con el suplemento de acetato. La biodegradación es
    posible en condiciones anaeróbicas. La bioconcentración en los peces
    de agua dulce es baja. La depuración es rápida.

         Las estimaciones de la exposición media calculadas a partir de
    diversos medios indican que la población en general está expuesta al
    cloroformo principalmente a través de los alimentos, el agua de
    bebida y el aire de los interiores en cantidades aproximadamente
    equivalentes. La inhalación estimada por conducto del aire exterior
    es considerablemente menor. La ingesta media estimada total es de
    aproximadamente 2 µg/kg de peso corporal por día. Los datos
    disponibles también indican que el agua de uso doméstico contribuye
    considerablemente a los niveles de cloroformo en el aire de los
    interiores y a la exposición total. La ingesta total estimada de
    algunos individuos que viven en lugares con un abastecimiento de
    agua corriente con concentraciones relativamente elevadas de
    cloroformo asciende a 10 µg/kg de peso corporal por día.

         Los animales y los seres humanos absorben bien el cloroformo
    después de la administración por vía oral, pero la cinética de la

    absorción depende del vehículo suministrado. Tras la exposición por
    inhalación, los seres humanos absorben del 60 al 80% de la cantidad
    inhalada. Los factores principales que afectan a la cinética de la
    absorción del cloroformo después de la inhalación son su
    concentración y la capacidad metabólica específica de la especie.
    Los seres humanos y los animales lo absorben fácilmente a través de
    la piel y se ha demostrado que durante la ducha la absorción dérmica
    del cloroformo del agua es apreciable. La hidratación de la piel
    parece acelerar la absorción de cloroformo.

         El cloroformo se distribuye en todo el cuerpo. Los niveles
    tisulares más elevados se alcanzan en el tejido adiposo, la sangre,
    el hígado, los riñones, los pulmones y el sistema nervioso. La
    distribución depende de la vía de exposición; los tejidos
    extrahepáticos reciben una dosis más elevada del cloroformo inhalado
    o absorbido por la piel que del cloroformo ingerido. Se ha
    demostrado que en varias especies animales y en el ser humano el
    cloroformo se transfiere a través de la placenta. El cloroformo se
    elimina principalmente como dióxido de carbono exhalado. El
    cloroformo no metabolizado se mantiene más tiempo en el tejido
    adiposo que en cualquier otro tejido.

         El citocromo P-450 cataliza la biotransformación oxidativa del
    cloroformo en triclorometanol. La pérdida de HCl del triclorometanol
    produce fosgeno como reactivo intermedio. El fosgeno puede
    destoxificarse por reacción con el agua produciendo dióxido de
    carbono o con tioles, inclusive con glutatión o cisteína,
    produciendo aductos. La reacción del fosgeno con proteínas tisulares
    está asociada con daño y necrosis celulares. Se observa un escaso
    enlace de los metabolitos del cloroformo con el ADN. El cloroformo
    también es objeto de una biotransformación reductiva catalizada por
    el P-450 que produce radicales de diclorometilo y éstos contraen
    enlaces covalentes con los lípidos tisulares. No se ha determinado
    el papel de la biotransformación reductiva en la citotoxicidad del
    cloroformo.

         Los animales y los seres humanos expuestos al cloroformo
    eliminan con el aire espirado el dióxido de carbono y el cloroformo
    que no se ha transformado. La fracción de dosis eliminada como
    dióxido de carbono varía según la dosis y la especie. La tasa de
    biotransformación en dióxido de carbono es más elevada en los
    microsomas hepáticos y renales de roedores (hámster, ratón, rata)
    que en los microsomas hepáticos y renales humanos. Además, el
    cloroformo se biotransforma más rápidamente en los microsomas
    renales del ratón que en los de la rata.

         El hígado es el órgano vulnerable a la toxicidad aguda en las
    ratas y en varias estirpes de ratones. La lesión hepática se
    caracteriza principalmente por una infiltración grasa temprana y
    células con forma de globo y evoluciona hacia la necrosis

    centrilobular seguida de necrosis general. El riñón es el órgano
    vulnerable en los ratones macho de otras estirpes más sensibles. La
    lesión renal comienza con una degeneración hidrópica que avanza
    hacia la necrosis de los tubos proximales. No se ha observado una
    toxicidad renal apreciable en las ratas hembra de ninguna estirpe.

         La toxicidad aguda varía según la raza, el sexo y el vehículo.
    En el ratón, la DL50 por vía oral oscila entre 36 y 1366 mg de
    cloroformo/kg de peso corporal, mientras que en las ratas oscila
    entre 450 y 2000 mg de cloroformo/kg de peso corporal. Después de
    una sola exposición de cuatro horas por inhalación, se observó
    toxicidad hepática en ratones y ratas cuando el nivel de cloroformo
    alcanzaba, respectivamente, 490 y 1410 mg/m3.

         Los efectos tóxicos del cloroformo más generales observados
    consisten en lesiones hepáticas. La gravedad de esos efectos por
    unidad de dosis administrada depende de la especie, del vehículo de
    administración y del método por el cual se haya administrado el
    cloroformo. La dosis más baja causante de lesión hepática observada
    es de 15 mg/kg de peso corporal por día, administrada a perros
    pachones en una base de pasta dentífrica durante un periodo de 7,5
    años. No se han examinado efectos con dosis más bajas. Se necesitan
    dosis algo más elevadas para producir efectos hepatotóxicos en otras
    especies. En esos estudios, aunque la duración de la exposición
    variaba, los niveles sin efectos adversos observados oscilaban entre
    15 y 125 mg/kg de peso corporal por día.

         Se han observado efectos en el riñón de ratones macho de
    estirpes sensibles y en la rata F-344. Se han observado efectos
    graves en una estirpe especialmente sensible de ratones macho con
    dosis de sólo 36 mg/kg de peso corporal por día.

         La inhalación de cloroformo seis horas por día durante siete
    días consecutivos produjo atrofia de las glándulas de Bowman y
    neoplasia ósea en la concha nasal de ratas F-344. El nivel en que no
    se observaron esos efectos fue de 14,7 mg/m3 (3 ppm). La
    importancia de dichos efectos se está investigando más a fondo en
    estudios de larga duración.

         El cloroformo administrado por sonda en un vehículo de aceite
    de maíz en dosis de 138 a 477 mg/kg de peso corporal por día indujo
    tumores hepáticos en ratones. Sin embargo, dosis semejantes de
    cloroformo administradas en el agua de bebida no produjeron tumores
    hepáticos en ratones. Por otra parte, en estudios de
    iniciación/promoción, el cloroformo administrado en el agua de
    bebida como promotor parecía inhibir el desarrollo de tumores
    hepáticos iniciados por dietilnitrosamina en ratones. Así pues, el
    vehículo y/o el método de administración del cloroformo es una
    variable importante en relación con la inducción de tumores
    hepáticos en el ratón.

         El cloroformo administrado por sonda en aceite de maíz en dosis
    de 90 a 200 mg/kg de peso corporal por día indujo tumores renales en
    ratas. Sin embargo, en esa especie se observaron efectos semejantes
    tras la administración de cloroformo en el agua de bebida, lo que
    indica que la reacción no depende exclusivamente del vehículo
    utilizado.

         Los efectos carcinogénicos del cloroformo en el hígado y los
    riñones de roedores parecen estar estrechamente relacionados con
    efectos citotóxicos y de replicación celular observados en los
    órganos vulnerables. Se ha observado que los efectos en la
    replicación celular eran paralelos a las modificaciones de las
    respuestas carcinogénicas al cloroformo inducidas por el vehículo y
    por la modalidad de administración. Las observaciones realizadas
    indican que el cloroformo tiene poca o ninguna capacidad para
    inducir mutaciones genéticas o daños directos de otro tipo en el
    ADN. Por otra parte, el cloroformo no parece poder iniciar tumores
    hepáticos en ratones ni de inducir síntesis imprevistas de ADN  in
     vivo. Por otra parte, el cloroformo puede promover la neoplasia
    hepática cuando se administra en un vehículo oleoso. Por
    consiguiente, es probable que, tras la administración prolongada de
    cloroformo, la citotoxicidad seguida de proliferación celular sea la
    causa más importante del desarrollo de tumores hepáticos y renales
    en los roedores.

         Algunos datos limitados sugieren que el cloroformo es tóxico
    para el feto, pero sólo en dosis tóxicas para la madre.

         En general, el cloroformo provoca en el ser humano los mismos
    síntomas de toxicidad que en los animales. En el ser humano, la
    anestesia puede causar la muerte por arritmia e insuficiencia
    respiratoria y cardíaca. En el ser humano también se ha observado
    necrosis de los tubos renales y disfunción renal. Los niveles más
    bajos en los que se haya comunicado toxicidad hepática debida a la
    exposición ocupacional al cloroformo se sitúan entre 80 y 160
    mg/m3 (con un periodo de exposición inferior a cuatro meses) en un
    estudio y entre 10 y 1000 mg/m3 (con periodos de exposición de uno
    a cuatro años) en otro estudio. La dosis letal media de un adulto se
    estima en unos 45 g, pero hay grandes diferencias de vulnerabilidad
    de un individuo a otro. En algunos estudios epidemiológicos hay
    ciertas indicaciones de que existe una asociación entre la
    exposición a los subproductos de la desinfección del agua de bebida
    y el cáncer colorrectal y de vejiga. Sin embargo, hay factores
    confusos y otras insuficiencias que ponen en entredicho esos
    estudios. Las pruebas de la carcinogenicidad del agua de bebida
    clorada en el ser humano son insuficientes. Además, los subproductos
    de la desinfección no pueden atribuirse al cloroformo por sí solo.

         El cloroformo es tóxico en las fases embriolarvales de algunas
    especies de anfibios y de peces. La CL50 más baja comunicada es de

    0,3 mg/litro en las fases embriolarvales de  Hyla crucifer. El
    cloroformo es menos tóxico para los peces y para  Daphnia magna. La
    CL50 de varias especies de peces se halla entre 18 y 191 mg/litro.
    Hay pocas diferencias de sensibilidad entre los peces de agua dulce
    y salada. La CL50 más baja comunicada en  Daphnia magna es de 29
    mg/litro. El cloroformo es poco tóxico para las algas y otros
    microorganismos.

         El Grupo Especial llegó a la conclusión de que los datos
    disponibles son suficientes para fijar una ingesta diaria tolerable
    sin efectos neoplásicos e ingestas con riesgos carcinogénicos
    específicos del cloroformo sobre la base de los estudios realizados
    en especies animales; las dosis servirán como orientación para que
    las autoridades competentes fijen límites de exposición. Sin
    embargo, se advierte que, cuando las circunstancias locales exijan
    optar entre el cumplimiento de límites microbiológicos y el de
    límites para subproductos de la desinfección tales como el
    cloroformo, debe siempre prevalecer la calidad microbiológica.
     Nunca debe comprometerse una desinfección eficaz.

         Sobre la base del estudio de Heywood et al. (1979) en el cual
    se observó una ligera hepatotoxicidad (aumento de las enzimas del
    suero hepático y quistes grasos) en perros pachones que habían
    ingerido 15 mg/kg de peso corporal por día en pasta dentífrica
    durante 7,5 años, incorporando un factor de incertidumbre de 1000
    (x10 para la variación entre especies, x10 para la variación dentro
    de la especie y x10 para utilizar un nivel con efectos en lugar de
    sin efectos y un estudio subcrónico), se obtiene una ingesta diaria
    tolerable (IDT) de 15 µg/kg de peso corporal por día.

         En función de los datos disponibles sobre los mecanismos
    determinantes, el método que se considera más apropiado para
    establecer orientaciones fundadas en los tumores hepáticos de
    ratones es dividir un nivel sin efectos de proliferación celular por
    un factor de incertidumbre. A partir del nivel sin efectos
    observados de citoletalidad y proliferación celular en ratones
    B6C3F1, de 10 mg/kg de peso corporal por día administrados en
    aceite de maíz durante tres semanas, comunicado en el estudio de
    Larson et al. (1993b), incorporando un factor de incertidumbre de
    1000 (x10 para la variación entre especies, x10 para la variación
    dentro de la especie y x10 para la gravedad del efecto, es decir,
    carcinogenicidad, y estudio subcrónico) se obtiene una IDT de 10
    µg/kg de peso corporal por día.

         Se reconoce que los tumores renales en ratas también pueden
    estar asociados con letalidad y proliferación celular. Sin embargo,
    dado que no se dispone de datos sobre proliferación celular en la
    estirpe en la que se observaron tumores y la información sobre
    proliferación y letalidad celulares es de corto plazo (una sola
    sonda y exposición por inhalación durante siete días), se considera

    prematuro alejarse del modelo establecido por defecto (es decir,
    fases múltiples linearizadas) como base para estimar el riesgo de
    cáncer durante una vida. La ingesta diaria total que se considera
    asociada con un riesgo excesivo de 10-5 durante una vida, sobre la
    base de la inducción de tumores renales (adenomas y adenocarcinomas)
    en ratas macho en el estudio de Jorgenson et al. (1985), es de 8,2
    µg/kg de peso corporal por día.

         Los niveles de cloroformo en las aguas superficiales son
    generalmente bajos y no se prevé que constituyan un peligro para los
    organismos acuáticos. Sin embargo, niveles más elevados de
    cloroformo en las aguas superficiales como consecuencia de las
    descargas o los derrames industriales tal vez sean peligrosos en las
    fases embriolarvales de algunas especies acuáticas.


See Also:
        Chloroform (CHEMINFO)
        Chloroform (IARC Summary & Evaluation, Supplement 7, 1987)
        Chloroform (IARC Summary & Evaluation, Volume 1, 1972)
        Chloroform (IARC Summary & Evaluation, Volume 20, 1979)
        Chloroform (ICSC)
        Chloroform (PIM 121)