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
joint venture of the United Nations Environment Programme, the
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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).
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 distr