Chloroform Detection and Quantification


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Chloroform displayed.svg
Chloroform 3D.svg
IUPAC name
Systematic name Trichloromethane
Other names Formyl trichloride, Methane trichloride, Methyl trichloride, Methenyl trichloride, TCM, Freon 20, R-20, UN 1888
CAS number 67-66-3 Yes check.svgY
PubChem 6212
EC number 200-663-8
KEGG C13827
ChEBI 35255
RTECS number FS9100000
ChemSpider ID 5977
Molecular formula CHCl3
Molar mass 119.38 g/mol
Appearance Colorless liquid
Density 1.483 g/cm3
Melting point -63.5 °C
Boiling point 61.2 °C
Solubility in water 0.8 g/100 ml (20 °C)
Refractive index (nD) 1.4459
Molecular shape Tetrahedral
MSDS External MSDS
R-phrases R22, R38, R40, R48/20/22
S-phrases (S2), S36/37
NFPA 704
NFPA 704.svg
Flash point Non-flammable
U.S. Permissible
exposure limit (PEL)
50 ppm (240 mg/m3) (OSHA)
Supplementary data page
Structure and
n, εr, etc.
Phase behaviour
Solid, liquid, gas
Spectral data UV, IR, NMR, MS
Yes check.svg(what is this?) (verify)
Except where noted otherwise, data are given for materials in their standard state (at 25 °C, 100 kPa)
Infobox references

Chloroform is the organic compound with formula CHCl3. This colorless, sweet-smelling, dense liquid is a trihalomethane. It is also considered somewhat hazardous. Several million tons were produced annually as a precursor to teflon and refrigerants, but the latter application is being phased out.[1]




[edit] Production

In industry, chloroform is produced by heating a mixture of chlorine and either chloromethane or methane.[1] At 400-500 °C, a free radical halogenation occurs, converting the methane or chloromethane to progressively more chlorinated compounds:

CH4 + Cl2 → CH3Cl + HCl
CH3Cl + Cl2CH2Cl2 + HCl
CH2Cl2 + Cl2 → CHCl3 + HCl

Chloroform undergoes further chlorination to give CCl4:

CHCl3 + Cl2 → CCl4 + HCl

The output of this process is a mixture of the four chloromethanes, chloromethane, dichloromethane, chloroform (trichloromethane), and carbon tetrachloride, which are then separated by distillation.[1]

[edit] Deuterochloroform

An archaic industrial route to chloroform involved the reaction of acetone (or ethanol) with sodium hypochlorite or calcium hypochlorite, known as the haloform reaction.[1] The chloroform can be removed from the coproducts by distillation. This reaction is still used for the production of bromoform and iodoform. The haloform process is obsolete for the production of ordinary chloroform. It is, however, used to produce CDCl3.[citation needed] Deuterochloroform can also be prepared by the reaction of sodium deuteroxide with chloral hydrate,[citation needed] or from ordinary chloroform.[2]

[edit] Inadvertent synthesis of chloroform

The haloform reaction can also occur inadvertently in domestic settings. Sodium hypochlorite solution (chlorine bleach) mixed with common household liquids such as acetone, butanone, ethanol, or isopropyl alcohol may produce some chloroform, in addition to other compounds such as chloroacetone, or dichloroacetone.

[edit] Uses

The major use of chloroform today is in the production of the chlorodifluoromethane (R-22), a major precursor to tetrafluoroethylene:

CHCl3 + 2 HF → CHClF2 + 2 HCl

The reaction is conducted in the presence of a catalytic amount of antimony pentafluoride. Chlorodifluoromethane is then converted into tetrafluoroethylene, the main precursor to Teflon. Before the Montreal Protocol, chlorodifluoromethane (R22) was also popular refrigerant.

[edit] As a solvent

Chloroform is a common solvent in the laboratory because it is relatively unreactive, miscible with most organic liquids, and conveniently volatile. Chloroform is used as a solvent in the pharmaceutical industry and for producing dyes and pesticides. Chloroform is an effective solvent for alkaloids in their base form and thus plant material is commonly extracted with chloroform for pharmaceutical processing. For example, it is used in commerce to extract morphine from poppies and scopolamine from Datura plants. Chloroform containing deuterium (heavy hydrogen), CDCl3, is a common solvent used in NMR spectroscopy. It can be used to bond pieces of acrylic glass (also known under the trade names Perspex and Plexiglas). A solvent of phenol:chloroform:isoamyl alcohol 25:24:1 is used to dissolve non-nucleic acid biomolecules in DNA and RNA extractions.

[edit] As a reagent in organic synthesis

As a reagent, chloroform serves as a source of the dichlorocarbene CCl2 group.[3] It reacts with aqueous sodium hydroxide usually in the presence of a phase transfer catalyst to produce dichlorocarbene, CCl2.[4][5] This reagent effects ortho-formylation of activated aromatic rings such as phenols, producing aryl aldehydes in a reaction known as the Reimer-Tiemann reaction. Alternatively the carbene can be trapped by an alkene to form a cyclopropane derivative.

[edit] History

Chloroform was discovered in July, 1831 by the American physician Samuel Guthrie,[6] and independently a few months later by the French chemist Eugène Soubeiran[7] and Justus von Liebig[8] in Germany, all of them using variations of the haloform reaction. Soubeiran produced chloroform through the action of chlorine bleach powder (calcium hypochlorite) on acetone (2-propanone) as well as ethanol. Chloroform was named and chemically characterised in 1834 by Jean-Baptiste Dumas.[9]

Chloroform in its liquid state shown in a test tube

Chloroform was developed on 4 November 1847 by James Young Simpson, head of midwifery at Edinburgh Hospital/University, and was mainly used as an anesthetic. Inhaling chloroform vapors depresses the central nervous system of a patient, causing dizziness, fatigue and unconsciousness[citation needed], allowing a doctor to perform simple surgery or various, otherwise painful, operations. In 1847, the Edinburgh obstetrician James Young Simpson first used chloroform for general anesthesia during childbirth. The use of chloroform during surgery expanded rapidly thereafter in Europe. In the United States, chloroform began to replace ether as an anesthetic at the beginning of the 20th century; however, it was quickly abandoned in favour of ether upon discovery of its toxicity, especially its tendency to cause fatal cardiac arrhythmia analogous to what is now termed “sudden sniffer’s death“. Ether is still the preferred anesthetic in some developing nations due to its high therapeutic index (~1.5-2.2) [10] and low price.

One possible mechanism of action for chloroform is that it increases movement of potassium ions through certain types of potassium channels in nerve cells.[11]

[edit] Safety

As might be expected for an anesthetic, inhaling chloroform vapors depresses the central nervous system. It is immediately dangerous to life and health at approximately 500 ppm, according to the United States National Institute for Occupational Safety and Health. Breathing about 900 ppm for a short time can cause dizziness, fatigue, and headache. Chronic chloroform exposure may cause damage to the liver (where chloroform is metabolized to phosgene) and to the kidneys, and some people develop sores when the skin is immersed in chloroform.

Though normally minute and diluted amounts are exposed to humans, chloroform can be harmful. Chloroform can be exposed via inhalation, ingestion of drinking water and foods made with chlorinated water, as well as dermal contact. It is readily excreted through exhalation, and minuscule amounts are excreted through urination and feces. The Environmental Protection Agency reports, a fatal oral dose of chloroform may be as low as 10 mL (14.8 g), with death due to respiratory or cardiac arrest [12].

Animal studies have shown that miscarriages occur in rats and mice that have breathed air containing 30 to 300 ppm of chloroform during pregnancy and also in rats that have ingested chloroform during pregnancy. Offspring of rats and mice that breathed chloroform during pregnancy have a higher incidence of birth defects, and abnormal sperm have been found in male mice that have breathed air containing 400 ppm chloroform for a few days. The effect of chloroform on reproduction in humans is unknown.

Chloroform once appeared in toothpastes, cough syrups, ointments, and other pharmaceuticals, but it has been banned as a consumer product in the United States since 1976.[13]

The National Toxicology Program’s eleventh report on carcinogens[14] implicates it as reasonably anticipated to be a human carcinogen, a designation equivalent to International Agency for Research on Cancer class 2A. It has been most readily associated with hepatocellular carcinoma.[15][16] Caution is mandated during its handling in order to minimize unnecessary exposure; safer alternatives, such as dichloromethane, have resulted in a substantial reduction of its use as a solvent.

During prolonged storage hazardous amounts of phosgene can accumulate in the presence of oxygen and ultraviolet light. To prevent accidents, commercial chloroform is stabilized with ethanol or amylene, but samples that have been recovered or dried no longer contain any stabilizer, and caution must be taken. Suspicious bottles should be tested for phosgene. Filter-paper strips, moistened with 5% diphenylamine, 5% dimethylaminobenzaldehyde, and then dried, turn yellow in phosgene vapor.

[edit] References

  1. ^ a b c d M. Rossberg et al. “Chlorinated Hydrocarbons” in Ullmann’s Encyclopedia of Industrial Chemistry 2006, Wiley-VCH, Weinheim. doi:10.1002/14356007.a06_233.pub2
  2. ^ Canadian Patent 1085423
  3. ^ Srebnik, M.; Laloë, E. “Chloroform” Encyclopedia of Reagents for Organic Synthesis” 2001 John Wiley.doi:10.1002/047084289X.rc105
  4. ^1,6-Methano[10]annulene“, Org. Synth., 1988, ; Coll. Vol. 6: 731
  5. ^ Gokel, G. W.; Widera, R. P.; Weber, W. P. (1988), “Phase-Transfer Hofmann Carbylamine Reaction: tert-Butyl Isocyanide“, Org. Synth., ; Coll. Vol. 6: 232
  6. ^ Samuel Guthrie (1832). “.”. Am. J. Sci. And Arts 21: 64.
  7. ^ Eugène Soubeiran (1831). “.”. Ann. Chim. 48: 131.
  8. ^ Justus Liebig (1832). “Ueber die Verbindungen, welche durch die Einwirkung des Chlors auf Alkohol, Aether, ölbildendes Gas und Essiggeist entstehen”. Annalen der Pharmacie 1 (2): 182–230. doi:10.1002/jlac.18320010203.
  9. ^ Jean-Baptiste Dumas (1834). “Untersuchung über die Wirkung des Chlors auf den Alkohol”. Annalen der Pharmacie 107 (41): 650–656. doi:10.1002/andp.18341074103.
  10. ^ Calderone, F.A. J. Pharmacology Experimental Therapeutics, 1935, 55(1), 24-39,
  11. ^ Patel, Amanda J.; Honoré, Eric; Lesage, Florian; Fink, Michel; Romey, Georges; Lazdunski, Michel (May 1999). written at Valbonne, France. “Inhalational anesthetics activate two-pore-domain background K+ channels“. Nature Neuroscience 2 (5): 422–426. doi:10.1038/8084. ISSN 1097-6256. PMID 10321245.
  12. ^
  13. ^The National Toxicology Program: Substance Profiles: Chloroform CAS No. 67-66-3” (pdf). Retrieved 2007-11-02.
  14. ^11th Report on Carcinogens“. Retrieved 2007-11-02.
  15. ^Centers for Disease Control and Prevention: CURRENT INTELLIGENCE BULLETIN 9“.
  16. ^National Toxicology Program: Report on the carcinogenesis bioassay of chloroform“.

[edit] External links

EPA Guidelines

Chloroform (CASRN 67-66-3)

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Main Contents

Reference Dose for Chronic Oral Exposure (RfD)
Reference Concentration for Chronic Inhalation Exposure
Carcinogenicity Assessment for Lifetime Exposure
Revision History


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CASRN 67-66-3

Health assessment information on a chemical substance is included in IRIS
only after a comprehensive review of chronic toxicity data by U.S. EPA
health scientists from several Program Offices and the Office of Research
and Development. The summaries presented in Sections I and II represent
a consensus reached in the review process. Background information and
explanations of the methods used to derive the values given in IRIS are
provided in the Background Documents.


File First On-Line 01/31/1987

Category (section)
Last Revised
Oral RfD Assessment
on-line 10/19/01
Inhalation RfC Assessment
Carcinogenicity Assessment
on-line 10/19/01

Chronic Health Hazard Assessments for Noncarcinogenic Effects

Reference Dose for Chronic Oral Exposure (RfD)

CASRN — 67-66-3
Last Revised — 10/19/01

The oral Reference Dose (RfD) is based on the assumption that thresholds
exist for certain toxic effects such as cellular necrosis. It is expressed
in units of mg/kg/day. In general, the RfD is an estimate (with uncertainty
spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without
an appreciable risk of deleterious effects during a lifetime. Please refer
to the Background Document for an elaboration of these concepts. RfDs
can also be derived for the noncarcinogenic health effects of substances
that are also carcinogens. Therefore, it is essential to refer to other
sources of information concerning the carcinogenicity of this substance.
If the U.S. EPA has evaluated this substance for potential human carcinogenicity,
a summary of that evaluation will be contained in Section II of this file.

Oral RfD Summary

Traditional Approach

For comparison purposes, an RfD was developed using the
traditional NOAEL/LOAEL approach. The results of this method are provided
below. This is the same approach and RfD result reported on IRIS (01/13/87).

Critical Effect
Experimental Doses*
Moderate/marked fatty
cyst formation in the
liver and elevated SGPT

Dog, chronic oral bioassay

Heywood et al., 1979

NOAEL: none

LOAEL: 15 mg/kg/day
(converted to 12.9 mg/kg/day)


*Conversion Factors and Assumptions — 15 mg/kg/day × 6 days/7 days = 12.9 mg/kg/day.

__I.A.2. Principal and Supporting Studies (Oral RfD)

Heywood, R; Sortwell, RJ; Noel, PRB; et
al. (1979) Safety evaluation of toothpaste containing chloroform: III. Long-term
study in beagle dogs. J Environ Pathol Toxicol 2:835-851.

Heywood et al. (1979) exposed groups of eight male and eight female
beagle dogs to doses of 15 or 30 mg chloroform/kg/day. The chemical was
given orally in a toothpaste base in gelatin capsules, 6 days/week for
7.5 years. This was followed by a 20- to 24-week recovery period. Eight
dogs of each sex served as an untreated group and a final group of 16
dogs (8/sex) received an alternative nonchloroform toothpaste (vehicle
control). Four male dogs (one each from the low- and high-dose chloroform
groups, the vehicle control group, and the untreated control group) and
seven female dogs (four from the vehicle control group and three from
the untreated control group) died during the study. In the low-dose group,
levels of serum glutamate-pyruvate transaminase (SGPT, also known as alanine
aminotransferase) were increased by an average of about 40% compared with
control, with the effects being statistically significant from week 130
through week 364. In the high-dose group, SGPT levels tended to average
about twice those in the control group, and the differences were statistically
significant from week 6 throughout treatment. After 14 weeks of recovery,
SGPT levels remained significantly increased in the high-dose group, but
not in the low-dose group, when compared with the controls. After 19 weeks
of recovery, SGPT levels were not significantly increased in either treated
group when compared with the controls. The authors concluded that the
increases in SGPT levels were likely the result of minimal liver damage.
Serum alkaline phosphatase (SAP) and SGPT levels were also moderately
increased (not statistically significant) in the treated dogs at the end
of the treatment period when compared with the controls. Microscopic examinations
were conducted on the major organs. The most prominent microscopic effect
observed in the liver was the presence of “fatty cysts,” which were described
as aggregations of vacuolated histiocytes. The fatty cysts were observed
in the control and treated dogs, but were larger and more numerous (i.e.,
higher incidence of cysts rated as “moderate or marked,” as opposed to
“occasional or minimal”) in the treated dogs than in the control dogs
at both doses. The prevalence of moderate or marked fatty cysts was 1/27
in control animals, 9/15 in low dose animals, and 13/15 in high dose animals.
Nodules of altered hepatocytes were observed in both treated and control
animals, and therefore were not considered related to treatment. No other
treatment-related nonneoplastic or neoplastic lesions were reported for
the liver, gall bladder, cardiovascular system, reproductive system, or
urinary system. A NOAEL was not identified in this study. However, a LOAEL
of 15 mg/kg/day was identified, based on elevated SGPT levels and increased
incidence and severity of fatty cysts (U.S. EPA, 1998a).

Benchmark Dose (BMD) Approach

Selection of Data Sets for Modeling

The following data sets were selected for BMD modeling:

  • Incidence of fatty cysts in liver and SGPT levels of dogs (Heywood
    et al., 1979)
  • Histological evidence of renal cytotoxicity in male rats exposed via
    drinking water (Hard et al., 2000)
  • Increased labeling index in kidney of female mice exposed via drinking
    water (Larson et al., 1994b)
  • Increased labeling index in liver of female rats exposed via gavage
    in corn oil (Larson et al., 1995b)

    These studies were chosen because
    they all provide quantitative dose-response data for sensitive indicators
    of chloroform toxicity.

BMD Modeling of Selected Data Sets

The detailed results of the BMD model fitting are presented in Appendix
B of the Toxicological Review of Chloroform. Within a data set, the preferred
model was selected based on the quality of the model fit to the data.

As seen, the kidney LI data set from Larson et al. (1994b) could not
be adequately described by any of the continuous models. This is because
even though the response was statistically significant, the magnitude
of the response was small in comparison to normal variability, and the
data did not form a smooth dose-response relationship (tending to first
increase and then decrease as dose increased). The liver and kidney LI
data sets from Larson et al. (1995b) were reasonably well fit by the Hill
equation, with BMD values of 64-75 mg/kg/day. However, the software was
not able to estimate a benchmark dose limit (BMDL) value in either case.
The data sets from the studies by Hard et al. (2000) and by Heywood et
al. (1979) were adequately fit by one or more of the dichotomous models,
with the best fit being given by the log-logistic and the quantal-linear
models, respectively. The preferred BMD of 70 mg/kg/day based on the renal
cytotoxicity data of Hard et al. (2000) is similar to the BMD values derived
for the LI data from Larson et al (1995b), but is significantly higher
than the preferred BMD based on the incidence of fatty cysts in dogs (1.7
mg/kg/day) reported by Heywood et al. (1979). The basis for this marked
difference in BMD between studies is not known, but the data suggest that
liver toxicity in the dog is a more sensitive endpoint of chloroform toxicity
than renal or liver cytotoxicity in rodents.

Calculation of the BMD-Based RfD

Critical Effect
Experimental Doses*
Moderate/marked fatty
cyst formation in the
liver and elevated SGPT

Dog, chronic oral bioassay

Heywood et al., 1979

BMDL10 : 1.2
(converted to 1.0 mg/kg/day)

The BMDL10 provided in the table represents
the 95% confidence lower bound on the dose associated with a 10% extra
risk based on the prevalence of animals with moderate to marked fatty
cysts in liver and elevated SGPT. The value of the BMDL10
was calculated from the data of Heywood et al. (1979) using EPA’s BMDS
software Version 1.2. The value derived from the BMD modeling (1.2 mg/kg/day)
was adjusted by a factor of 6/7 to account for exposure 6 days per week.

Uncertainty and Modifying Factors (Oral RfD)

UF = 100

In the benchmark dose approach, an uncertainty factor (UF) of 10 was
used to account for interspecies extrapolation, and a UF of 10 was used
to protect sensitive subpopulations. In the NOAEL/LOAEL approach, an additional
factor of 10 was used to account for extrapolation from a LOAEL to a NOAEL
(total UF = 1,000). No additional factors were required to account for
extrapolation from short term to long term (the study duration was 7.5
years) or to account for limitations in the database.

MF = 1

No additional modifying factors (MFs) were considered necessary because
there are no substantial concerns or limitations in the derivation of
the RfD that are not accounted for in the UFs described above.

Additional Studies/Comments (Oral RfD)

In general, the NOAEL/LOAEL approach for
derivation of an RfD is subject to a number of limitations, most of which
are addressed by use of the BMD approach (U.S. EPA, 1995). Thus, the RfD
based on the BMD approach is generally preferred, unless there are insufficient
dose-response data to support derivation of a reliable BMD.

In this case, the dose-response data set from the critical study (Heywood
et al., 1979) is composed of only two doses plus a control group. This
is considered to be a limitation, as the shape of the dose-response curve
is difficult to define with only three values, especially when the lowest
dose yields a response that is well above the benchmark response. Nevertheless,
the data do yield curve fits of adequate quality, so the results of the
BMD approach are considered preferable to the NOAEL/LOAEL approach.

Note that, in this particular case, the two approaches (NOAEL/LOAEL
and benchmark) yield equal RfD values. This is consistent, albeit coincidental,
with the results from the default LOAEL/NOAEL method.

Many other studies in animals support the conclusion that the liver
and/or the kidney are the key target organs for chloroform-induced toxicity.
Most of these studies have been performed in rats and mice, and most yield
LOAEL values that are substantially higher than those observed in dogs.

In a study conducted by Palmer et al. (1979), in which rats were administered
daily oral doses of 60 mg chloroform/kg/day in a toothpaste vehicle, treatment-related
effects included a decrease in plasma but not erythrocyte, cholinesterase
in females, a decrease in liver weight in females, and a marginal but
consistent and progressive retardation in weight gain in both sexes. The
authors stated that although minor histological changes in the liver were
noted, there was no evidence of severe fatty infiltration, fibrosis, or
bile duct abnormalities in the livers of treated animals. The authors
concluded that there was no evidence of treatment-related toxic effects
in the liver. However, the “minor histopathological” changes in the liver
were not described and the presence of any fatty infiltration that would
be designated as less than severe was not reported. Therefore, these results
could not be compared to those reported in the dog study. The LOAEL for
this study was 60 mg/kg/day.

A slight (2%-3% vs. 7%-8%) increase in moderate to severe fatty degeneration
of the liver was seen in ICI mice given 60 mg but not 17 mg chloroform/kg/day
in a toothpaste vehicle for 80 weeks (Roe et al., 1979). However, no effects
were evident when the incidences of fatty and nonfatty liver degeneration
were combined in the ICI or three other mice strains. No other noncancer
effects attributable to chloroform were noted. A NOAEL of 17 mg/kg/day
and a LOAEL of 60 mg/kg/day were identified from this study.

No treatment-related noncancer effects were noted in rats administered
chloroform in drinking water for 23 months at time-weighted average doses
up to 160 mg/kg/day (Jorgenson et al., 1982, 1985). However, subsequent
review of the histopathology slides from this study revealed evidence
that chloroform produced a moderate to low level of renal proximal tubule
injury associated with cell turnover indicative of cytotoxicity (Hard
et al., 2000). These changes were noted in the high-dose (160 mg/kg) group
males as early as 12 months but were increased in grade by 18 months.
Similar changes were found in the mid-dose males (81 mg/kg), although
at a lower grade, in the 18-month and 2-year dose groups. These changes
were not seen in controls or the low-dose group. Therefore, the identified
NOAEL for noncancer effects for this study is 38 mg chloroform/kg/day,
with the LOAEL at 81 mg/kg/day.

In mice exposed to chloroform in drinking water, mortality within the
first 3 weeks was significantly increased in the two highest dose groups,
130 and 263 mg/kg/day, but was comparable with controls after that time
(Jorgenson et al., 1982). Early mortality and behavioral effects (e.g.,
lassitude, lack of vigor) were apparently related to reduced water consumption
among some treated mice in the two highest dose groups. A significant
increase in liver fat in mice was noted at doses of 65 mg/kg/day and higher
at 3 months, but only at doses of 130 and 263 mg/kg/day by 6 months. Liver
fat content was not reported for any later time points or at terminal
sacrifice; therefore, the relevance of this observation as an adverse
effect rather than an adaptive response could not be assessed. No increased
incidence of liver tumors was reported, and the presence or absence of
nonneoplastic histopathological alterations was not described. These data
indicate that doses of 130 to 263 mg/kg/day may produce adverse effects
in mice; however, these effects may be secondary to decreased water consumption.

Reproductive/developmental toxicity studies were also considered in
the selection of the critical study/effect for the reference dose in the
event the fetus represented a more sensitive population. These included
studies in rats (Thompson et al., 1974), in rabbits (Thompson et al.,
1974), and in mice (NTP, 1988). In the developmental studies in rabbits
and rats, no treatment-related effects were noted when chloroform was
administered by gavage in corn oil during gestation at doses of 50 mg/kg/day
or less (Thompson et al., 1974). In the rabbit study, a clear dose-response
was absent and the effects noted in offspring of dams administered chloroform
at doses up to 50 mg/kg/day (the highest dose tested) on days 6 to 18
of gestation were not considered to be treatment-related (Thompson et
al., 1974). In rats, the only effect noted was a significant reduction
in fetal weight found only in offspring of dams given chloroform at the
highest dose tested, 126 mg/kg/day, on days 6 to 15 of gestation (Thompson
et al., 1974). No fetal effects attributed to chloroform treatment were
noted in this rat study for the lower dose groups (up to 50 mg/kg/day
during gestation). A NOAEL of 50 mg/kg/day was identified for both studies.

In a two-generation reproductive study in mice, no significant effects
were seen in any reproductive parameter assessed in either the parental
or the F1 generations at doses up to
41 mg/kg/day administered by gavage in corn oil (NTP, 1988). Systemic
toxicity was not evaluated in the parental generation. However, increased
liver weights and liver lesions, described as mild to moderate degeneration
of centrilobular hepatocytes accompanied by single-cell necrosis, were
noted in F1 females, but not males,
exposed both in utero and postnatally at a dose of 41 mg/kg/day. Postnatal
exposure in the F1 generation began
at postnatal day 22 and continued until the birth of the F2
generation (mice were mated at 64 to 84 days of age). The F1
offspring in the two lower dose groups, 6.6 and 16 mg/kg/day, were not
evaluated histopathologically; therefore, no NOAEL or LOAEL could be definitively
established for this study. A dose of 41 mg/kg/day may represent the LOAEL;
however, the amount of in utero exposure was not estimated, nor was the
contribution of in utero exposure to liver toxicity assessed. Because
quantitative data were available only for the control and high-dose groups,
the study was not selected for benchmark modeling.

In the reproductive/developmental studies, both maternal toxicity and
effects on the fetus or offspring occurred at doses higher than those
that produced evidence of liver toxicity in the dog study. Therefore,
these were not used as the critical study for derivation of the RfD.
For more detail on Susceptible Populations,
exit to the toxicological review,
Section 4.7

Confidence in the Oral RfD

Study — Medium
Database — Medium
RfD — Medium

The overall confidence in this RfD assessment
is medium. The database on noncancer effects in animals is extensive,
and data are adequate to derive reliable dose-response curves for key
endpoints. Confidence is not rated higher because data in humans are limited,
and extrapolation from animals to humans (with an attendant uncertainty
factor of 10) is required.

For more detail on Characterization of Hazard and
Dose Response, exit to the
toxicological review, Section 6

EPA Documentation and Review of the Oral RfD

Source Document — U.S. EPA, 2001

This assessment was peer reviewed by external scientists. Their comments
have been evaluated carefully and incorporated in finalization of this
IRIS Summary. A record of these comments is included as an appendix to
U.S. EPA (2001). To
review this appendix, exit to the toxicological review, Appendix A, External
Peer Review — Summary of Comments and Disposition (PDF)

Other EPA Documentation — U.S. EPA, 1994, 1997, 1998a-c, 2001

Agency Consensus Date — 7/27/2001

EPA Contacts (Oral RfD)

Please contact the IRIS Hotline for
all questions concerning this assessment or IRIS, in general, at (202)566-1676
(phone), (202)566-1749 (FAX) or
(internet address).

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Reference Concentration for Chronic Inhalation Exposure (RfC)

(Not available. To be developed)

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Carcinogenicity Assessment for Lifetime Exposure

CASRN — 67-66-3
Last Revised — 10/19/01

Section II provides information on three
aspects of the carcinogenic assessment for the substance in question;
the weight-of-evidence judgment of the likelihood that the substance is
a human carcinogen, and quantitative estimates of risk from oral exposure
and from inhalation exposure. The quantitative risk estimates are presented
in three ways. The slope factor is the result of application of a low-dose
extrapolation procedure and is presented as the risk per (mg/kg)/day.
The unit risk is the quantitative estimate in terms of either risk per
µg/L drinking water or risk per µg/cu.m air breathed. The third
form in which risk is presented is a concentration of the chemical in
drinking water or air associated with cancer risks of 1 in 10,000, 1 in
100,000, or 1 in 1,000,000. The rationale and methods used to develop
the carcinogenicity information in IRIS are described in The Risk Assessment
Guidelines of 1986 (EPA/600/8-87/045) and in the IRIS Background Document.
IRIS summaries developed since the publication of EPA’s more recent Proposed
Guidelines for Carcinogen Risk Assessment also utilize those Guidelines
where indicated (Federal Register 61(79):17960-18011, April 23, 1996).
Users are referred to Section I of this IRIS file for information on long-term
toxic effects other than carcinogenicity.

Evidence for Human Carcinogenicity

Weight-of-Evidence Characterization

Under the 1986 U.S. EPA Guidelines for
Carcinogen Risk Assessment, chloroform has been classified as Group B2,
probable human carcinogen, based on “sufficient evidence” of carcinogenicity
in animals (U.S. EPA, 1998a).

Under the Proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA,
1996; U.S. EPA, 1999), chloroform is likely to be carcinogenic to humans
by all routes of exposure
under high-exposure conditions that lead
to cytotoxicity and regenerative hyperplasia in susceptible tissues (U.S.
EPA, 1998a,b). Chloroform is not likely to be carcinogenic to humans
by any route of exposure
under exposure conditions that do not cause
cytotoxicity and cell regeneration. This weight-of-evidence conclusion
is based on: 1) observations in animals exposed by both oral and inhalation
pathways which indicate that sustained or repeated cytotoxicity with secondary
regenerative hyperplasia precedes, and is probably required for, hepatic
and renal neoplasia; 2) there are no epidemiological data specific to
chloroform and, at most, equivocal epidemiological data related to drinking
water exposures that cannot necessarily be attributed to chloroform amongst
multiple other disinfection byproducts; and 3) genotoxicity data on chloroform
are essentially negative, although there are some scattered positive results
that generally have limitations such as excessively high dose or with
confounding factors. Thus, the weight-of-evidence of the genotoxicity
data on chloroform supports a conclusion that chloroform is not strongly
mutagenic, and that genotoxicity is not likely to be the predominant mode
of action underlying the carcinogenic potential of chloroform. Although
no cancer data exist for exposures via the dermal pathway, the weight-of-evidence
conclusion is considered to be applicable to this pathway as well, because
chloroform absorbed through the skin and into the blood is expected to
be metabolized and to cause toxicity in much the same way as chloroform
absorbed by other exposure routes.

For more detail on Characterization of Hazard and
Dose Response, exit to the
toxicological review, Section 6

For more detail on Susceptible Populations, exit
to the toxicological review,
Section 4.7

Human Carcinogenicity Data

Inadequate. There are no epidemiological data attributing cancer
to exposure to chloroform per se. Although there are some equivocal
epidemiological data relating a weak association of drinking water exposures
to bladder, rectal and colon cancer (Morris et al. 1992 ; McGeehin et
al., 1993; Vena et al. 1993; Morris, 1995; King and Marrett, 1996; Doyle
et al., 1997; Freedman et al., 1997; Cantor et al, 1998; Hildesheim et
al., 1998), these studies can not attribute to chloroform among multiple
other disinfection byproducts (DBPs) (SAB, 2000, ATSDR, 1997; IPCS, 2000).
Morris et al. (1992) did a meta-analysis that pooled the relative risks
from ten cancer epidemiology studies in which there was a presumed exposure
to chlorinated water and its byproducts and estimated that approximately
10,000 cases of rectal and bladder cancer cases per year could be associated
with exposure to DBPs in chlorinated water in the United States. Later,
Poole (1997) reviewed the studies available to Morris et al. (1992) plus
three additional studies (McGeehin et al., 1993; Vena et al., 1993; and
King and Marrett, 1996). Poole (1997) observed that there was considerable
heterogeneity among the data and that there was evidence of publication
bias within the body of literature. In addition, Poole found that the
aggregate estimates reported by Morris et al. were sensitive to small
changes in the analysis (e.g., addition or deletion of a single study).
Based on the observations, Poole recommended that the cancer epidemiology
data considered in the Morris evaluation should not be combined into a
single summary estimate and that the data had limited utility for risk
assessment purposes. Based on the available cancer epidemiology database,
bladder cancer studies provide the strongest evidence for an association
between exposure to chlorinated water and cancer. Based on the studies
of Cantor et al. (1985), McGeehin et al. (1993), King and Marrett (1996),
Freedman et al. (1997), and Cantor et al. (1998), EPA calculated that
the population attributable risk (the fraction of a disease that could
be eliminated if the exposure of concern were eliminated) for bladder
cancer ranged from 2% to 17% (U.S. EPA, 1998c). However, these calculations
are based on a number of assumptions, including the assumption that there
is a cause-effect relationship between exposure to chlorinated drinking
water and increased risk of bladder cancer. This assumption is subject
to considerable uncertainty, especially because findings are not consistent
within or between studies. Evaluation of these studies by application
of standard criteria for establishing causality from epidemiological observations
(strength of association, consistency of findings, specificity of association,
temporal sequence, dose-response relation, biological plausibility) has
led EPA to conclude that the current data are insufficient to establish
a causal relationship between exposure to chloroform and increased risk
of cancer (U.S. EPA, 1998a). Moreover, if, in the future, the weight-of-evidence
does reach a point where a causal link is established between exposure
to chlorinated water and increased risk of bladder or other types of cancer,
it could not be concluded from epidemiological studies of this type that
chloroform per se is carcinogenic in humans, as chlorinated water contains
numerous disinfection byproducts besides chloroform that are potentially
carcinogenic (U.S. EPA, 1998a).

Animal Carcinogenicity Data

Adequate. At high doses, chloroform
has been reported to be carcinogenic in several chronic animal bioassays,
with significant increases in the incidence of liver tumors in male and
female mice and significant increases in the incidence of kidney tumors
in male rats and mice (U.S. EPA, 1994, 1998c). When examining the biology
of the tumor production, the occurrence of tumors is demonstrably species-,
strain-, and gender-specific, and has only been observed under dose conditions
that caused cytotoxicity and regenerative cell proliferation in the target

In a gavage bioassay (NCI, 1976), Osborne-Mendel rats and B6C3F1 mice
were treated with chloroform in corn oil 5 times/week for 78 weeks (50
animals per sex per dose group). Male rats received 90 or 125 mg/kg/day;
females initially were treated with 125 or 250 mg/kg/day for 22 weeks
and 90 or 180 mg/kg/day thereafter. A decrease in survival rate and weight
gain was evident for all treated rats. A significant increase in kidney
epithelial tumors was observed in male rats (0% in controls, 8% in the
low dose and 24% in the high dose groups). Male mice received 100 or 200
mg/kg/day, raised to 150 or 300 mg/kg/day at 18 weeks; females were dosed
with 200 or 400 mg/kg/day, raised to 250 or 500 mg/kg/day. Survival rates
and weight gains were comparable for all groups except high dose female
mice which had a decreased survival. In mice, highly significant increases
in hepatocellular carcinomas were observed in both sexes (98% and 95%
for males and females at the high dose; 36% and 80% for males and females
at the low dose as compared with 6% of both matched and colony control
males , 0% in matched control females and 1% in colony control females).
Nodular hyperplasia of the liver was observed in many low dose male mice
that had not developed hepatocellular carcinoma. Hepatomas have also developed
in female strain A mice and NLC mice gavaged with chloroform (Eschenbrenner
and Miller, 1945; Rudali, 1967).

Jorgenson et al. (1985) administered chloroform (pesticide quality and
distilled) in drinking water to male Osborne-Mendel rats and female B6C3F1
mice at concentrations of 200, 400, 900, and 1,800 mg/L for 104 weeks.
These concentrations were reported by the author to correspond to 19,
38, 81, and 160 mg/kg/day for rats and 34, 65, 130, and 263 mg/kg/day
for mice. The combined benign and malignant renal tumor incidence in male
rats was 2%, 2%, 2%, 5%, 6% and 14% for the control, matched control,
19, 38, 81, and 160 mg/kg/day groups, respectively. A significant increase
in renal tumors (14%) in rats was observed in the highest dose group (160
mg/kg/day). A reevaluation of the histopathology of the slides (Hard et
al., 2000), found evidence of persistent cytotoxicity and regenerative
hyperplasia in all rats of the highest dose group. Similar changes were
also observed in rats at 81 mg/kg/day, but at a much lower incidence and
grade. Thus, the histopathology reexamination provides evidence supporting
chronic renal tubule injury as the mode of action underlying the renal
tumor response. The liver tumor incidence in female mice was not significantly

Chloroform administered in toothpaste was not carcinogenic to male C57B1,
CBA, CF-1, or female ICI mice or to beagle dogs. Male ICI mice administered
60 mg/kg/day were found to have an increased incidence of kidney epithelial
tumors (Roe et al., 1979; Heywood et al., 1979). A pulmonary tumor bioassay
in strain A/St mice was negative, as was one in which newborn C57X DBA2/F1
mice were treated s.c. on days 1 to 8 of life (Theiss et al., 1977; Roe
et al., 1968).

Matsushima (1994) exposed F344 rats (50/sex/group) and BDF1 mice (50/sex/group)
to chloroform vapor 6 hours/day, 5 days/week for 104 weeks. Rats were
exposed to concentrations of 0, 10, 30, or 90 ppm, and mice were exposed
to 0, 5, 30, or 90 ppm. In order to avoid short- term lethality, mice
in the two highest groups (30 and 90 ppm) were initially exposed to a
lower levels for 2-6 weeks before the long-term exposure. The time-weighted
average (TWA) for the 30 ppm group was 29.1 ppm and for the 90 ppm group
was 85.7 ppm (U.S. EPA, 1998a). Statistically significant increases in
the incidence of overall renal cell adenoma and renal cell carcinoma were
observed in male mice in the 30 (7/50) and 90 (12/48) ppm groups, when
compared to controls (0/50). The overall incidence rates of renal cell
carcinoma were statistically significantly increased in males in the 90-ppm
group (11/48) when compared to controls (0/50). There were no statistically
significant findings reported for female mice in any exposure groups.

Supporting Data for Carcinogenicity


Many studies have investigated the mutagenic potential of chloroform.
However, there are several reasons these studies must be reviewed carefully
and interpreted cautiously. For example, chloroform is relatively volatile,
so test systems not designed to prevent chloroform escape to the air may
yield unreliable results. Earlier studies in which appropriate P450-based
metabolic activation systems were absent are also likely to be unreliable.
Further, some older studies that used ethanol as a solvent or preservative
for chloroform may be confounded by formation of ethyl or diethyl carbonate,
which are potent alkylating agents. Another important issue is that studies
that focused on clastogenicity endpoints using excessively high doses
may be confounded by severe cytotoxicity, causing lysosomal or other releases
(Brusick, 1986).

In Vitro Studies

Two investigators reported DNA binding in studies with calf thymus DNA
in the presence of exogenous activation (DiRenzo et al., 1982; Colacci
et al., 1991). The study by DiRenzo et al. (1982) used ethanol as a solvent,
suggesting that ethyl carbonate formation might be a problem. In the study
by Colacci et al. (1991), addition of SKF-525A inhibited DNA binding,
suggesting that binding was mediated by a cytochrome P-450 mediated pathway,
as would be expected for chloroform. In interpreting these studies, it
is important to remember that cell-free systems may not always be a good
model for intact cellular processes.

Gene mutation studies in Salmonella typhimurium and E. coli
(Ames assay), including tests done under conditions designed to reduce
evaporation, are mostly negative, with or without activation with microsomes
from liver or kidney of rats or mice (Rapson et al., 1980; San Agustin
and Lim-Sylianco, 1978; Van Abbe et al., 1982; Uehleke et al., 1977; Gocke
et al., 1981; Roland-Arjona et al., 1991; Le Curieux et al., 1995; Kirkland
et al., 1981; Simmon et al., 1977). However, four studies have showed
positive results in bacteria. Varma et al. (1988) reported that chloroform
caused mutagenicity in five strains of S. typhimurium, but the
response was noted only at the lowest dose tested, and all higher doses
were not different from control. This unusual pattern casts some doubt
on these results. San Agustin and Lim-Sylianco (1997) reported that chloroform
caused DNA damage in Bacillus subtilis, and Wecher and Scher (1982)
reported that chloroform caused mutations in Photobacterium phosphoreum.
However, neither study reported the exposure concentrations that caused
these effects, so the relevance of these reports is uncertain. In addition,
the studies by Varma et al. (1988) and Wecher and Scher (1982) each used
ethanol as a diluent, raising the possibility that the positive effect
might be related to ethyl carbonate formation rather than to chloroform.
The majority of results reported for S. typhimurium and E. coli
exposed to the vapor phase were also negative (Van Abbe et al., 1982;
Pegram et al., 1997; Simmon, 1977; Sasaki et al., 1998). Pegram et al.
(1997) reported that chloroform was weakly positive at vapor concentrations
greater than 19,200 ppm (about 770 mg/L in the aqueous phase). Employing
physiologically based pharmacokinetic models, the authors estimated the
oral doses needed to produce the effect would exceed 2,000 mg/kg (approximately
twice the LD50).

Tests of genotoxicity are also mainly negative in fungi (Gualandi, 1984;
Mehta and von Borstel, 1981; Kassinova et al., 1981; Jagannath et al.,
1981). However, chloroform was shown to induce intrachromosomal recombination
in Saccharomyces cerevisiae at concentrations of 6,400 mg/L (Callen
et al., 1980) or 750 mg/L (Brennan and Schiestl, 1998). In the Brennan
and Schiestl study, addition of N-acetylcysteine reduced chloroform-induced
toxicity and recombination, suggesting a free radical may have been involved.
Chromosome malsegregation was also reported in Aspergillus nidulans
(Crebelli et al., 1988), but only at concentrations above 1,600 mg/L.
In all three of these positive studies, doses that caused positive results
also caused cell death, indicating that exposures were directly toxic
to the test cells.

Studies in intact mammalian cells are mainly negative (Larson et al.,
1994a; Perocco and Prodi, 1981; Butterworth et al., 1989; Kirkland et
al., 1981; White et al., 1979; Sturrock, 1977), although positive results
have been reported in a few systems. Increased sister chromatid exchange
was reported in human lymphocytes at a concentration of about 1,200 mg/L
without exogenous activation (Morimoto and Koizumi, 1983), and at a lower
concentration (12 mg/L) with exogenous activation (Sobti, 1984). In the
study by Sobti, the increase was quite small (less than 50%), and there
was an increase in the number of cells that did not exclude dye. This
suggests that the exposure levels that caused the mutagenic effect may
have been directly toxic to the cells. In addition, ethanol was used as
a dose vehicle. Mitchell et al. (1988) did not detect an increase in mutation
in mouse lymphoma cells at an exposure level of 2,100 mg/L in the absence
of exogenous activation, but did detect an effect at a concentration of
59 mg/L with exogenous activation.

In Vivo Studies

A number of different endpoints of chloroform genotoxicity have been
measured in intact animals exposed to chloroform either orally or by inhalation.
In studies of DNA binding in liver and kidney of mice and rats, negative
results have been reported at doses of 742 mg/kg, 119 mg/kg, and 48 mg/kg
(Diaz-Gomez and Castro, 1980; Reitz et al., 1982; Pereira et al., 1982).
However, positive results have been reported at doses as low as 2.9 mg/kg
(Colacci et al., 1991). But, in the study by Colacci et al. (1991), no
significant difference in binding was noted between multiple tissues (liver,
kidney, lung, and stomach), and there was no increase in binding with
phenobarital pretreatment. This suggests the binding may not have been
related to chloroform metabolism.

Studies based on signs of DNA damage or repair have been uniformly negative
(Larson et al., 1994a; Potter et al., 1996; Reitz et al., 1982; Mirsalis
et al., 1982). However, studies based on various signs of chromosomal
abnormalities have been mixed, with some studies reporting negative findings
at doses of 371 mg/kg and 800 mg/kg (Shelby and Witt, 1995; Topham, 1980),
while other studies report positive results at doses as low as 1.2 mg/kg
(Fujie et al., 1990). However, the positive result at low dose in the
study by Fujie et al. (1990) was observed following intraperitoneal exposure,
and positive results following oral exposure were not observed until a
dose level of 119 mg/kg. Morimoto and Koizumi (1983) observed an increase
in the frequency of sister chromatid exchange in bone marrow cells at
a dose of 50 mg/kg/day, but at 200 mg/kg/day, all of the mice died. As
discussed before, mutagenicity results observed following highly toxic
doses may have been confounded by cytotoxic responses and should be viewed
as being of uncertain relevance.

Several studies have reported negative findings for the micronucleus
test in rats and mice (Gocke et al., 1981; Salamone et al., 1981; Le Curieux,
1995), but several other studies have detected positive results, mainly
at exposure levels of 400-600 mg/kg (San Agustin and Lim-Sylianco, 1982;
Robbiano et al., 1998; Sasaki et al., 1998; Shelby and Witt, 1995). This
suggests that chloroform may be clastogenic, but it is important to note
that these doses are well above the level that causes cytotoxicity in
liver and kidney in most oral exposure studies in rodents.

Butterworth et al. (1998) did not detect an increase in mutation frequency
in male mice exposed by inhalation at an exposure level of 90 ppm, even
though this exposure did cause an increase in tumors in the study by Nagano
et al. (1998). Increased incidence of sperm head abnormalities was reported
in mice exposed at 400 ppm (Land et al., 1981), but was not observed in
mice exposed to 371 mg/kg intraperitoneally (Topham, 1980).

In Drosophila melanogaster larvae exposed to chloroform vapor,
gene mutation (Gocke et al., 1981) and mitotic recombination tests (Vogel
and Nivard, 1993) were both negative. Grasshopper embryos (Melanoplus
) did not display mitotic arrest at vapor concentrations
of 30,000 ppm, but an effect was seen at 150,000 ppm (Liang et al., 1983).
San Agustin and Lim-Syllianco (1981) reported a single positive and negative
result for host- mediated mutagenicity in Salmonella typhimurium,
but exposure levels were not reported in either case.

On the basis of the in vitro and in vivo studies reviewed above, even
though a role of mutagenicity cannot be completely ruled out, the majority
of available studies are negative, and many of the positive studies have
limitations (excessive doses or other confounding factors). Thus, the weight-of-evidence
supports the conclusion that chloroform is not strongly mutagenic, and
that genotoxicity is not likely to be the predominant mode of action underlying
the carcinogenic potential of chloroform. This conclusion is supported
by a number of other groups who have reviewed and evaluated the available
data on chloroform genotoxicity, including the International Commission
for Protection against Environmental Mutagens and Carcinogens (Lohman
et al., 1992), ILSI (1997), Health Canada (2000), and WHO (1998).

Mode of Action

1. Summary of Postulated Mode of Action

Studies in animals reveal that chloroform can cause an increased incidence
of kidney tumors in male rats and an increased incidence of liver tumors
in male and female mice. Available data suggest that tumors are produced
only at dose levels that result in cytotoxicity. These induced tumor responses
are postulated to be secondary to sustained or repeated cytotoxicity and
secondary regenerative hyperplasia. Chloroform’s carcinogenic effects
in rodent liver and kidney are attributed to oxidative metabolism-mediated
cytotoxicity in the target organs. Although chloroform undergoes both
oxidative and reductive cytochrome P450-mediated metabolism, it is the
oxidative (CYP2E1) metabolic pathway that predominates at low chloroform
exposures. This oxidative pathway produces highly tissue-reactive metabolites
(in particular phosgene) that lead to tissue injury and cell death. It
is likely that the electrophilic metabolite phosgene causes cellular toxicity
by reaction with tissue proteins and cellular macromolecules as well as
phospholipids, glutathione, free cysteine, histidine, methionine, and
tyrosine. The liver and kidney tumors induced by chloroform depend on
persistent cytotoxic and regenerative cell proliferation responses. The
persistent cell proliferation presumably would lead to higher probabilities
of cell mutation and subsequent cancer. The weight of the evidence indicates
that a mutagenic mode of action via DNA reactivity is not a significant
component of the chloroform carcinogenic process.

2. Identification of key events

There are essentially three key steps in the sequence of events that
lead to chloroform-induced tumorigenesis in the liver and kidneys of rodents.
The first step is oxidative metabolism of chloroform in the target organs,
kidney and liver. Numerous binding and metabolism studies (as described
in ILSI, 1997, and U.S. EPA, 1998a) provide support that chloroform is
metabolized by the oxidative cytochrome P450 (CYP2E1) pathway. This conclusion
is supported by the study of Constan et al. (1999) in Sv/129 wild type,
Sv/129 CYP2E1 null, and B6C3F1 mice. In the wild type of each strain,
exposure to 90 ppm chloroform for 6 hours per day for 4 consecutive days
resulted in severe hepatic and renal lesions along with increased cell
proliferation. With the same exposure, neither the cytotoxicity nor cell
proliferation occurred in the CYP2E1 null mouse or in the wild type of
either strains treated with the P450 inhibitor ABT.

Available evidence indicates that metabolism by CYP2E1 predominates
at low exposures and is rate-limiting to chloroform’s carcinogenic potential.
Reductive metabolism, if it occurs, can lead to free radicals and tissue
damage, but this pathway is absent or minor under normal physiological
conditions. The next key step is the resultant cytotoxicity and cell death
caused by the oxidative metabolites (with phosgene as the significant
toxic intermediate). Regenerative cell proliferation follows the hepatotoxicity
and nephrotoxicity as measured by labeling index in mouse kidney and liver
and rat kidney from chloroform-treated animals. This increase in cell
division is responsible for the increased probability of cancer.

3. Strength, consistency, specificity of association

There are numerous cases where exposure to chloroform causes an increase
in cytotoxicity (as evidenced by histopathological evaluation and/or increased
labeling index) without any observable increase in cancer incidence. These
data indicate that chloroform exposures that are adequate to cause cytotoxicity
and regenerative cell proliferation do not always lead to cancer. However,
there are no cases where a tumorogenic response has been observed in which
evidence of cell regeneration is not also observed at the same or lower
dose as that which caused an increase in tumors. This consistency of evidence
(i.e., cell regeneration is detected in all cases of tumorigenicity) is
strong evidence supporting the conclusion that cell regeneration is a
mandatory precursor for tumorigenicity.

Evidence for a link between sustained cytotoxicity/regenerative hyperplasia
and cancer is strongest in the kidney. In male Osborne-Mendel rats exposed
to chloroform in water for 2 years (Jorgenson et al., 1985), a statistically
significant increase in renal tumors was observed at a concentration of
1,800 ppm (160 mg/kg/day). A re-analysis of the histopathological slides
from this study (Hard et al., 2000) revealed evidence for sustained cytotoxicity
and cell proliferation in the kidney at exposures of 900 ppm (81 mg/kg/day)
or higher. Likewise, in BDF1 mice exposed
to chloroform by inhalation at 5, 30, or 90 ppm for 6 hours/day, 5 days/week
(Nagano et al., 1998), increased incidence of renal tumors was observed
in male mice at the two higher doses, whereas females showed no significant
tumor response. Templin et al. (1998) duplicated this exposure regimen
in order to study whether the treatment caused cytotoxicity and regenerative
hyperplasia. These authors observed cytotoxicity and hyperplasia in the
kidneys of male mice exposed to 30 or 90 ppm throughout a 90-day exposure
period, but not in females. This observation is consistent with the hypothesis
that sustained cytotoxicity and regenerative hyperplasia are key events
in the neoplastic response of the kidney to chloroform.

Available data also indicate that cytotoxicity and regenerative hyperplasia
are required for liver cancer, although the strength of this conclusion
is somewhat limited because most of the observations are based on short-term
rather than long-term histological or labeling index measurements. For
example, in the B6C3F1 mouse, corn oil gavage (bolus dosing) at the same
doses that resulted in liver tumors in the study by NCI (1976) also caused
hepatic cytolethality and a cell proliferative response at both 4 days
and 3 weeks (Larson et al., 1994b,c). Similarly, exposure of female B6C3F1
mice to chloroform in drinking water at levels that did not induce liver
tumors (Jorgenson et al., 1985) also did not induce hepatic cytolethality
or cell proliferation at 4 days or 3 weeks (Larson et al., 1994b). This
consistency of the data (i.e., evidence of cytolethality and/or regenerative
hyperplasia is always observed in cases of increased liver tumors) supports
the conclusion that this liver cancer also occurs via a mode of action
involving regenerative hyperplasia.

4. Dose-response relationship

Chloroform-induced liver tumors in mice are only seen after bolus corn
oil dosing. Mouse liver tumors are not found following administration
by other routes (drinking water and inhalation). Rat liver tumors are
not induced by chloroform following either drinking water or corn oil
gavage administration. Kidney tumors are found in mice exposed to chloroform
via inhalation or toothpaste preparations, and in rats when exposed via
drinking water or corn oil gavage. Kidney and liver tumors develop only
at doses that cause persistent cytotoxicity and regenerative proliferation,
regardless of route of exposure or dosing regime. The overall dose-response
for the cytotoxicity and cell proliferation responses is nonlinear. All
key events and tumor effects depend on the dose-rate as shown by the difference
in oil gavage versus drinking water administration (ILSI, 1997; U.S. EPA,

5. Temporal relationship

As noted above, there is very strong evidence from short-term and long-term
histological and labeling index studies in mice and rats that cytotoxicity
and cell proliferation always precede the occurrence of increased kidney
or liver tumor effects in long-term bioassays. For example, a re-evaluation
of serial sacrifice data from the chloroform 2-year drinking water bioassay
in Osborne-Mendel rats revealed a linkage between toxicity in the renal
tubules and tumor development and showed that renal toxicity preceded
tumor development (Hard and Wolf, 1999; Hard et al., 2000).

6. Biological plausibility and coherence

The theory that sustained cell proliferation to replace cells killed
by toxicity, viral, or other insults such as physical abrasion of tissues
can be a significant risk factor for cancer is plausible and generally
accepted (Correa, 1996). It is logical to deduce that sustained cytotoxicity
and regenerative cell proliferation may result in a greater likelihood
of mutations being perpetuated with the possibility of more of these resulting
in uncontrolled growth. It may also be that continuous stimulus of proliferation
by growth factors involved in inflammatory responses increases the probability
that damaged cells may slip through cell cycle check points carrying DNA
alterations that would otherwise be repaired. Current views of cancer
processes support both these possibilities. There are no data on chloroform
that allow the events that occur during cell proliferation to be directly
observed. A high proliferation rate alone is not assumed to cause cancer;
tissues with naturally high rates of turnover do not necessarily have
high rates of cancer and tissue toxicity in animal studies does not invariably
lead to cancer. Nevertheless, regenerative proliferation associated with
persistent cytotoxicity appears to be a risk factor of consequence.

7. Role of genotoxicity

As noted above, the question whether chloroform or a metabolite is mutagenic
has been tested extensively across different phylogenetic orders (i.e.,
bacterial, eukaryotic, and mammalian systems). Predominately negative
results are reported in all test systems, with no pattern of mutagenicity
seen in any one system considered to be a competent predictor. Positive
results appear sporadically in the database, but they generally have problems
with high dose or other confounding issues. ILSI (1997) considered results
from 40 tests by the quantitative weight-of-evidence method for heterogeneous
genetic toxicology databases from the International Commission for Protection
against Environmental Mutagens and Carcinogens (ICEMC) (Lohman et al.,
1992). This method scores relative DNA reactivity, with a maximum positive
score being +100 and maximum negative -100. The maximum positive score
obtained among 100 chemical databases has been +49.7 (triazaquone) and
the maximum negative has been -27.7. The score for chloroform was -14.3.

Testing of chloroform in the p53 heterozygous knockout mouse shows no
tumor effect (Gollapudi et al., 1999). Heterozygous p53 males were dosed
up to 140 mg/kg and females up to 240 mg/kg via corn oil gavage for 13
weeks. This model is known to respond most effectively to mutagenic carcinogens.

Products of oxidative and reductive metabolism of chloroform are highly
reactive. Such species are unstable and will likely react with cytoplasmic
molecules before reaching nuclear DNA. Such reactive species (e.g., phosgene)
have not been evaluated separately for genetic toxicity, and because of
reactivity, would not be amenable to study and would not likely be able
to transport from the cellular site of production to the nucleus.

Comparative examination of both oxidative and reductive metabolism for
structural analogues and chloroform has revealed that carbon tetrachloride,
which is largely metabolized to a free radical via the reductive pathway,
results in cell toxicity, not mutagenicity. Moreover, chloroform and carbon
tetrachloride show very different patterns of liver toxicity (i.e., carbon
tetrachloride’s toxicity is more consistent with free radical production
and chloroform’s is not). For methylene chloride, glutathione conjugation
results in mutagenic metabolites. When rat glutathione transferase gene
copies are introduced into Salmonella, bromodichloromethane produces
mutagenic metabolites; the fact that chloroform in this system did so
only marginally and only at high toxic doses (Pegram et al., 1997) supports
a conclusion that the reductive pathway does not contribute to chloroform’s
toxicity or carcinogenicity.

In initiation-promotion studies, chloroform at the highest test dose
of the drinking water bioassay does not promote development of hepatic
lesions in rats or two strains of mice, nor does it initiate or act as
a cocarcinogen. Administered in oil, chloroform was a promoter in the
rat liver in initiation-promotion protocols. These results are more consistent
with the postulated mode of action than with any mutagenic potential.

8. Effects on children

The central questions asked in a mode of action analysis are, 1) whether
the standard assumption that a mode of action observed in animals is relevant
to humans holds true in a particular case, and 2) what the nature of the
mode of action implies about the shape of the dose response relationship.
In the case of chloroform the conclusions have been that the rodent mode
of action can be assumed to be relevant to humans and that a nonlinear
approach is most appropriate. The next question is whether the data lead
one to anticipate similarities or differences in response by sex or age.

Ideally, one would have adequate data to compare each of the key events
of chloroform toxicity and subsequent carcinogenicity in tissues of adults
with those of the developing fetus and young. This kind of information
is currently not to be found. In the absence of data on the fetus and
young specific to chloroform, an evaluation is made as to whether a cogent
biological rationale exists for determining that the postulated mode of
action is applicable to children (EPA, 1999). There is no suggestion from
available studies of chloroform to indicate that children or fetuses would
be qualitatively more sensitive to its effects than adults. The developing
organism would not be expected to be particularly sensitive to cytotoxic
agents at minimally toxic levels because cell division is proceeding rapidly
and repair capacity at the molecular and cellular level is high. This
is reflected by the relatively low incidence of spontaneous tumors in
developing and young organisms. Moreover, the reproductive and developmental
studies available, while they have limitations, show that fetal effects
are seen only at doses at which maternal toxicity is evident. Research
would be needed to further explore whether there are circumstances in
which this relationship does not hold. Research would also be needed to
discover whether there is some other mode of action, not seen in rodents,
that might be possible. Presently, there are no clues from in vivo or
in vitro studies as to what alternative mode of action might be considered.
In keeping with traditional toxicologic evaluations, chloroform has been
tested in lifetime studies with high level doses to provide maximal opportunities
for toxicologic effects to manifest themselves in multiple tissues and
organs through multiple mechanisms. In the absence of data to the contrary,
this approach is considered to provide evidence for lack of potential
for significant response, other than those noted, even for sensitive individuals
and life stages.

The mode of action analyzed as well as all other potential modes of
action identified required that chloroform be metabolized by cytochrome
P450 (CYP2E1) (SAB (2000), p.2). When this is considered along with the
comparison of this enzyme activity between adults and the young there
is confidence in assuming similarity in response among life stages. Further
research on the processes of cell injury, death and regeneration would
increase this confidence by addressing any uncertainty about potential
quantitative similarity. The literature does not reveal any such quantitative
data at present.

Given the above, it is reasonable to assume that: 1) The reactive metabolite
inside the cell should have similar effects by reacting with and disrupting
macromolecules in the cells of fetuses, children and adults, 2) Cell necrosis
and reparative replication are not likely to be qualitatively different
in various stages in life, 3) Cancer risk to the fetus or children would
be a function of cytotoxic injury, like in adults, and protecting these
life stages from sufficient cytotoxicity to elicit this response should
protect against cancer risks. Further research would be needed to assess
whether there are significant quantitative differences between life stages
which have not yet been elucidated.

It can be noted that if data indicated that it were appropriate to apply
a linear approach to part of a lifetime, such as the first 3 years of
life, the resulting risk would be represented by a small increment of
the total dose per body weight over a lifetime since most of a 70 year
life is at an adult body weight. When this total is divided by 70 years
to derive the lifetime average daily dose, the small increment of early
dose does not significantly increase risk.

9. Conclusion regarding cancer mode of action

The weight of the evidence supports the conclusion that chloroform-induced
tumors in liver and kidney are produced only at dose levels that result
in repeated or sustained cytotoxicity and regenerative cell proliferation.
A wide range of evidence across different species, sexes, and routes of
exposure implicates oxidative CYP2E1 metabolism leading to persistent
cytotoxicity and regenerative cell proliferation as events that precede
and are associated with tumor formation. The cytochrome P450 oxidative
metabolism that leads to oxidative damage and ensuing cell growth, involving
basic tissue responses to cellular toxicity and death, is common to humans
and rodents. No data exist indicating that the mode of action observed
in rodents is not also likely to apply to humans.

Available data on the mutagenic potential of chloroform are mixed, but
the majority of tests are negative, and some of the positive results are
observed only at extreme exposure conditions. Thus, the weight of the
evidence indicates that chloroform is not a strong mutagen and that neither
chloroform nor its metabolites readily bind to DNA. On the basis of these
results and the results of studies that evaluated other endpoints of mutagenicity,
it seems likely that even though a role for mutagenicity cannot be excluded
with certainty, chloroform does not produce carcinogenic effects primarily
by a specific genotoxic mechanism.

The proposed dose-response relationship for chloroform tumorigenesis
by the cytotoxicity-regenerative hyperplasia mode of action will be nonlinear,
as it is dependent on biochemical and histopathological events that are
nonlinear. The dose-response assessment would ideally be based on use
of phosgene dosimetry because it marks the rate-limiting step of oxidative
metabolism. The toxicokinetic modeling to support this phosgene approach
is not currently available, so the dose-response assessment is based on
the tumor precursor event of cytotoxicity to project a level of exposure
that will be protective against the key event of regenerative hyperplasia.

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Quantitative Estimate of Carcinogenic Risk from Oral Exposure

In accord with proposed EPA guidelines for cancer risk
assessment (U.S. EPA, 1996), the method used to characterize and quantify
cancer risk from a chemical depends on what is known about the mode of
action of carcinogenicity and the shape of the cancer dose-response curve
for that chemical. A default assumption of linearity is appropriate when
evidence supports a mode of action of gene mutation due to DNA reactivity,
or another mode of action that is anticipated to be linear. The linear
approach is used as a matter of policy if the mode of action of carcinogenicity
is not understood. Alternatively, an assumption of nonlinearity is appropriate
when there is no evidence for linearity and sufficient evidence to support
an assumption of nonlinearity. In this case, the carcinogenicity may be
a secondary effect of toxicity that itself is a threshold phenomenon (U.S.
EPA, 1996).

In the case of chloroform, the mode of action of carcinogenicity
is reasonably well understood. Available data indicate that chloroform
is not strongly mutagenic and chloroform is not expected to produce rodent
tumors via a mutagenic mode of action (ILSI, 1997). Rather, there is good
evidence that carcinogenic responses observed in animals are associated
with regenerative hyperplasia that occurs in response to cytolethality
(ILSI, 1997; U.S. EPA, 1998a,b). Because cytolethality occurs only at
exposure levels above some critical dose level, a nonlinear approach is
considered the most appropriate method for characterizing the cancer risk
from chloroform.

The Proposed Guidelines for Carcinogenic Risk Assessment
(U.S. EPA, 1996) state that when the mode-of-action analysis based on
available data indicates that “the carcinogenic response is secondary
to another toxicity that has a threshold, the margin-of-exposure analysis
performed for toxicity is the same as is done for a noncancer endpoint,
and an RfD for that toxicity may be considered in the cancer assessment.”
For chloroform, available evidence indicates that chloroform-induced carcinogenicity
is secondary to cytotoxicity and regenerative hyperplasia; hence, the
Agency relies on a nonlinear dose-response approach and the use of a margin-of-exposure
analysis for cancer risk. The Agency has also chosen not to rely on a
mathematical model to estimate a point of departure for cancer risk estimate,
because the mode of action indicates that cytotoxicity is the critical
effect and the reference dose value is considered protective for this

RfD and Margin of Exposure

For more discussion of margin of exposure (MOE), see
the Toxicological Review for Chloroform. Based on the kidney tumor of
the drinking water study (Jorgenson et al., 1985), a point of departure
(Pdp or LED10) of 23 mg/kg/day can be calculated using quantitative
modeling of tumor dose-response data. Comparing the Pdp to the RfD of
0.01 mg/kg/day leads to a MOE of 2,000, which is considered large. Thus,
in this case, the RfD for noncancer effect is also considered adequately
protective of public health for cancer effects by the oral route, on the
basis of the nonlinear dose response for chloroform and the mode of action
for both cancer and noncancer effects having a common link through cytotoxicity.

As discussed above, the RfD for noncancer effects is
derived from the most sensitive endpoint in the most sensitive species.
The RfD is based on fatty cysts formation (fat accumulation) in the liver
and elevation of SGPT in dogs (Heywood et al., 1979). Hepatic fat accumulation
and elevated SGPT are considered early signs of impaired liver function
resulting from chloroform-induced cytotoxicity. This effect occurs at
doses at or below those that cause increased labeling index, morphological
changes, or cellular necrosis, so protection against this effect is believed
to protect against cytolethality and regenerative hyperplasia. Accordingly,
the RfD of 0.01 mg/kg/day presented in Section I.A.1 can be considered
protective against increased risk of cancer.

Summary of Risk Estimates

A dose of 0.01 mg/kg/day (equal to the RfD) can be considered
protective against cancer risk

____II.B.1.1. Oral Slope Factor — Not applicable (see

____II.B.1.2. Drinking Water Unit Risk — Not applicable
(see text).

Dose-Response Data (Carcinogenicity, Oral Exposure)

Dose-response data used to derive the RfD for chloroform
are presented in Section I.A.2.

Additional Comments (Carcinogenicity, Oral Exposure)

Because chloroform is a volatile chemical,
exposure to chloroform in drinking water may occur not only via direct
ingestion, but also by inhalation of chloroform released from household
uses of water (showering, cooking, washing, etc.) into indoor air. Therefore,
assessments of cancer and noncancer health effects from chloroform in
water should account for exposures by all pathways, including oral, inhalation,
and dermal.

Discussion of Confidence (Carcinogenicity, Oral Exposure)

Confidence in the cancer assessment for
chloroform is rated as medium. This is based on a strong database in animals
that supports the conclusion that cancer does not occur without antecedent
cytotoxicity and regenerative hyperplasia, leading in turn to the conclusion
that cancer risk is negligible at doses that do not result in cytotoxicity.
Confidence in this conclusion is tempered by absence of direct studies in
humans and by the finding that there are some positive results in studies
on the mutagenicity of chloroform, even though the weight-of-evidence indicates
that chloroform is not a strong mutagen and that a mutagenic mode of action
is not likely to account for the cancer responses observed in animals.

EPA is currently revising its guidelines for cancer risk assessment.
Among other issues, EPA is looking closely at how to assess whether a
postulated mode of action in adults is applicable to children. When the
guidelines are final, EPA will consider their impact on existing health
assessments on IRIS.

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Quantitative Estimate of Carcinogenic Risk from Inhalation Exposure

NOTE: The following evaluation of cancer risk
from chloroform inhalation was developed in 1987 and does not incorporate
newer data or the 1996 or 1999 draft cancer assessment guidelines.
EPA is currently working to revise the assessment for inhalation exposure.

Summary of Risk Estimates

____II.C.1.1. Inhalation Unit Risk — 2.3E-5 per (ug/m3).

____II.C.1.2. Extrapolation Method — Linearized multistage procedure, extra risk.

Air Concentrations at Specified
Risk Levels:

Risk Level
E-4 (1 in 10,000) 4E+0 µg/m3
E-5 (1 in 100,000) 4E-1 µg/m3
E-6 (1 in 1,000,000) 4E-2 µg/m3

Dose-Response Data for Carcinogenicity, Inhalation Exposure

Tumor Type — hepatocellular carcinoma
Test Animals — mouse, B6C3F1, female
Route — oral, gavage
Reference — NCI, 1976

—————- Dose —————

Additional Comments (Carcinogenicity, Inhalation Exposure)

This inhalation quantitative risk estimate
is based on data from a gavage study. Above doses are TWA; body weights
at the end of the assay were 35 g for males and 28 g for females. Vehicle
control animals were run concurrently and housed with test animals. All
treated animals experienced decreased body weight gain. Survival was reduced
in high-dose males and in all treated females. Experimental data for this
compound support complete absorption of orally administered chloroform
under conditions of this assay. There are no apparent species differences
in this regard. Extrapolation of metabolism-dependent carcinogenic responses
from mice to humans on the basis of body surface area is supported by
experimental data. The incidence data for both male and female mice were
used to derive slope factors of 3.3E-2 and 2.0E-1 per (mg/kg)/day, respectively.
The unit risk was prepared by taking a geometric mean of the slope factor
and assuming 100% for low doses of chloroform in air. The unit risk should
not be used if the air concentration exceeds 400 µg/m3,
because above this concentration the unit risk may not be appropriate.

Discussion of Confidence (Carcinogenicity, Inhalation Exposure)

Adequate numbers of animals were treated
and observed. Risk estimates derived from male rat kidney tumor data (2.4E-2)
(NCI, 1976) and studies by Roe et al. (1979) (1.0E-1) are generally supportive
of the risk estimate.

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EPA Documentation, Review, and Contacts (Carcinogenicity Assessment)

EPA Documentation

Source Document — U.S. EPA, 2001 (oral carcinogenicity assessment);
U.S. EPA, 1985, 1987 (inhalation carcinogenicity assessment)

This assessment was peer reviewed by external scientists.
Their comments have been evaluated carefully and incorporated in finalization
of this IRIS Summary. A record of comments on the oral carcinogenicity
assessment is included in an appendix to U.S. EPA (2001).
To review this appendix, exit to the toxicological review, Appendix A,
External Peer Review — Summary of Comments and Disposition (PDF)

EPA Review (Carcinogenicity Assessment)

Agency Consensus Date (oral carcinogenicity
assessment) — 7/27/2001
Verification Date (inhalation carcinogenicity assessment) – 8/26/1987

EPA Contacts (Carcinogenicity Assessment)

Please contact the IRIS Hotline for
all questions concerning this assessment or IRIS, in general, at (202)566-1676
(phone), (202)566-1749 (FAX) or
(internet address).

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_IV.  [reserved]
_V.  [reserved]


CASRN — 67-66-3
Last Revised — 10/19/01

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Inhalation RfC References

(Not applicable.)

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Revision History

CASRN — 67-66-3

03/01/1988 I.A.1. Dose conversion clarified
03/01/1988 I.A.2. LOAEL and RfD in text
03/01/1988 I.A.4. Text revised
03/01/1988 I.A.5. Text revised
06/30/1988 II. Carcinogen summary
06/30/1988 I.A.7. Primary contact changed
10/01/1989 I.B. Inhalation RfD now
under review
06/01/1990 IV.A.1. Area code for EPA contact
06/01/1990 IV.F.1. EPA contact changed
01/01/1991 II. Text edited
01/01/1991 II.C.1. Inhalation slope factor
removed (global change)
02/01/1991 II.C.3. Information on extrapolation
process included
02/01/1991 II.C.4. Text edited
03/01/1991 II.D.3. Primary contact changed
01/01/1992 IV. Regulatory actions
04/01/1992 IV.A.1. CAA regulatory action
07/01/1992 I.A. Clarify Schwetz citation
07/01/1992 VI.C. Oral RfD references
07/01/1992 VI.C. Carcinogenicity assessment
references on-line
09/01/1992 I.A.7. Primary contact changed
08/01/1995 I.B. EPA’s RfD/RfC and CRAVE
workgroups were discontinued in May, 1995. Chemical substance reviews
that were not completed by September 1995 were taken out of IRIS review.
The IRIS Pilot Program replaced the workgroup functions beginning
in September, 1995.
04/01/1997 III., IV., V. Drinking Water Health
Advisories, EPA Regulatory Actions, and Supplementary Data were removed
from IRIS on or before April 1997. IRIS users were directed to the
appropriate EPA Program Offices for this information.
12/10/1998 I.B. This chemical is being
reassessed under the IRIS Program.
10/19/2001 I.A.,VI Oral RfD and references
10/19/2001 II.B.,VI Oral carcinogenicty
assessment and references updated
01/08/2002 II.C.1. Corrected typographical
error in units in inhalation unit risk.
03/26/2002 Tox. Review Corrected list of external peer reviewers.

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CASRN — 67-66-3
Last Revised — 10/19/01

  • 67-66-3
  • Chloroform
  • Formyl Trichloride
  • Freon 20
  • Methane Trichloride
  • Methane, Trichloro-
  • Methenyl Chloride
  • Methenyl Trichloride
  • Methyl Trichloride
  • NCI-CO2686
  • R-20
  • TCM
  • Trichloroform
  • Trichloromethane

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2 responses to this post.

  1. Thank you for every other wonderful article.

    The place else could anyone get that type of information in such an ideal approach of writing?
    I have a presentation next week, and I’m at the look for such info.


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