Toxicological
Profile
for
CHLOROFORM
o
Agency for Toxic Substances and Disease Registry
U.S. Public Health Service
0.
c'
3
-------
ATSDR/TP-88/09
TOXICOLOCICAL PROFILE FOR
CHLOROFORM
Date Published — January 1989
Prepared by:
Syracuse Research Corporation
Under Contract No. 68-C8-0004
for
Agency for Toxic Substances and Disease Registry (ATSDR)
U.S. Public Health Service
in collaboration with
U.S. Environmental Protection Agency (EPA)
Technical editing/document preparation by:
Oak Ridge National Laboratory
under
DOE Interagency Agreement No. 1857-B026-A1
-------
DISCLAIMER
Mention of company name or product does not constitute endorsement by
the Agency for Toxic Substances and Disease Registry.
-------
FOREWORD
The Superfund Amendments and Reauthorization Act of 1986 (Public
Law 99-499) extended and amended the Comprehensive Environmental
Response, Compensation, and Liability Act of 1980 (CERCLA or Superfund)
This public law (also known as SARA) directed the Agency for Toxic
Substances and Disease Registry (ATSDR) to prepare toxicological
profiles for hazardous substances which are most commonly found at
facilities on the CERCLA National Priorities List and which pose the
most significant potential threat to human health, as determined by
ATSDR and the Environmental Protection Agency (EPA). The list of the 100
most significant hazardous substances was published in the Federal
Register on April 17, 1987.
Section 110 (3) of SARA directs the Administrator of ATSDR to
prepare a toxicological profile for each substance on the list. Each
profile must include the following content:
"(A) An examination, summary, and interpretation of available
toxicological information and epidemiologic evaluations on a
hazardous substance in order to ascertain the levels of significant
human exposure for the substance and the associated acute,
subacute, and chronic health effects.
(B) A determination of whether adequate information on the health
effects of each substance is available or in the process of
development to determine levels of exposure which present a
significant risk to human health of acute, subacute, and chronic
health effects.
(C) Where appropriate, an identification of toxicological testing
needed to identify the types or levels of exposure that may present
significant risk of adverse health effects in humans."
This toxicological profile is prepared in accordance with
guidelines developed by ATSDR and EPA. The guidelines were published in
the Federal Register on April 17, 1987. Each profile will be revised and
republished as necessary, but no less often than every three years, as
required by SARA.
The ATSDR toxicological profile is intended to characterize
succinctly the toxicological and health effects information for the
hazardous substance being described. Each profile identifies and reviews
the key literature that describes a hazardous substance's toxicological
properties. Other literature is presented but described in less detail
than the key studies. The profile is not intended to be an exhaustive
document; however, more comprehensive sources of specialty information
are referenced.
111
-------
Foreword
Each toxicological profile begins with a public health statement.
which describes in nontechnical language a substance's relevant
toxicological properties. Following the statement is material that
presents levels of significant human exposure and, where known,
significant health effects. The adequacy of information to determine a
substance's health effects is described in a health effects summary.
Research gaps in toxicologic and health effects information are
described in the profile. Research gaps that are of significance to
protection of public health will be identified by ATSDR, the National
Toxicology Program of the Public Health Service, and EPA. The focus of
the profiles is on health and toxicological information; therefore, we
have included this information in the front of the document.
The principal audiences for the toxicological profiles are health
professionals at the federal, state, and local levels, interested
private sector organizations and groups, and members of the public. We
plan to revise these documents in response to public comments and as
additional data become available; therefore, we encourage comment that
will make the toxicological profile series of the greatest use.
This profile reflects our assessment of all relevant toxicological
testing and information that has been peer reviewed. It has been
reviewed by scientists from ATSDR, EPA, the Centers for Disease Control,
and the National Toxicology Program. It has also been reviewed by a
panel of nongovernment peer reviewers and was made available for public
review. Final responsibility for the contents and views expressed in
this toxicological profile resides with ATSDR.
James 0. Mason, H.D., Dr. P.H.
Assistant Surgeon General
Administrator, ATSDR
iv
-------
CONTENTS
FOREWORD
LIST OF FIGURES ix
LIST OF TABLES xt
1. PUBLIC HEALTH STATEMENT 1
1.1 WHAT IS CHLOROFORM? L
1.2 HOW MIGHT I BE EXPOSED TO CHLOROFORM? 1
1.3 HOW DOES CHLOROFORM GET INTO MY BODY? 2
1.4 HOW CAN CHLOROFORM AFFECT MY HEALTH? 2
1.5 IS THERE A MEDICAL TEST TO DETERMINE IF I HAVE BEEN
EXPOSED TO CHLOROFORM? 3
1.6 WHAT LEVELS OF EXPOSURE HAVE RESULTED IN HARMFUL
HEALTH EFFECTS? 3
1.7 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT
MADE TO PROTECT HUMAN HEALTH? 7
2. HEALTH EFFECTS SUMMARY 9
2.1 INTRODUCTION 9
2.2 LEVELS OF SIGNIFICANT EXPOSURE 10
2.2.1 Key Studies and Graphical Presentations 10
2.2.1.1 Inhalation LO
2.2.1.2 Oral 19
2.2.1.3 Dermal 23
2.2.2 Biological Monitoring as a Measure of
Exposure and Effects 23
2.2.3 Environmental Levels as Indicators of
Exposure and Effects 25
2.2.3.1 Levels found in the environment 25
2.2.3.2 Human exposure potential 26
2. 3 ADEQUACY OF DATABASE 26
2.3.1 Introduction 26
2.3.2 Health Effect End Points 27
2.3.2.1 Introduction and graphic summary 27
2.3.2.2 Descriptions of highlights of graphs .... 27
2.3.2.3 Summary of relevant ongoing research .... 30
2.3.3 Other Information Needed for Human
Health Assessment 31
2.3.3.1 Pharmacokinetics and mechanisms
of action 31
2.3.3.2 Monitoring of human biological samples .. 31
2.3.3.3 Environmental considerations 32
3. CHEMICAL AND PHYSICAL INFORMATION 33
3.1 CHEMICAL IDENTITY 33
3.2 PHYSICAL AND CHEMICAL PROPERTIES 33
-------
Concents
4. TOXICOLOGICAL DATA 37
4.1 OVERVIEW 37
4.2 TOXICOKINETICS '.'.'." 33
4.2.1 Absorption 33
4.2.1.1 Inhalation 38
4.2.1.2 Oral '. 39
4.2.1.3 Dermal 39
4.2.2 Distribution 40
4.2.2.1 Inhalation 40
4.2.2.2 Oral 41
4.2.2.3 Dermal 41
4.2.3 Metabolism 41
4.2.3.1 Inhalation 41
4.2.3.2 Oral 41
4.2.3.3 Dermal 44
4.2.4 Excretion 44
4.2.4.1 Inhalation 44
4.2.4.2 Oral 44
4.2.4.3 Dermal 45
4.3 TOXICITY 45
4.3.1 Lethality and Decreased Longevity 45
4.3.1.1 Inhalation 45
4.3.1.2 Oral 45
4.3.1.3 Dermal 46
4.3.2 Systemic/Target Organ Toxicity 46
4.3.2.1 Liver effects 46
4.3.2.2 Kidney effects 51
4.3.2.3 CNS effects 54
4.3.3 Developmental Toxicity 55
4.3.3.1 Inhalation 55
4.3.3.2 Oral 56
4.3.3.3 Dermal 57
4.3.3.4 General discussion 57
4.3.4 Reproductive Toxicity 57
4.3.4.1 Inhalation 57
4.3.4.2 Oral 58
4.3.4.3 Dermal 58
4.3.4.4 General discussion 58
4.3.5 Genotoxicity 58
4.3.5.1 Human 58
4.3.5.2 Nonhuman 58
4.3.5.3 General discussion 58
4.3.6 Carcinogenic ley ' 61
4.3.6.1 Inhalation 61
4.3.6.2 Oral 61
4.3.6.3 Dermal 64
4.3.6.4 General discussion 64
4.4 INTERACTIONS WITH OTHER CHEMICALS 68
vi
-------
Concents
5. MANUFACTURE, IMPORT, USE, AND DISPOSAL 71
5.1 OVERVIEW 71
5.2 PRODUCTION 71
5.3 IMPORT 71
5.4 USES 71
5.5 DISPOSAL 72
6. ENVIRONMENTAL FATE 73
6.1 OVERVIEW 73
6.2 RELEASES TO THE ENVIRONMENT 73
6.3 ENVIRONMENTAL FATE 73
6.3.1 Air 73
6.3.2 Water 75
6.3.3 Soil 76
7. POTENTIAL FOR HUMAN EXPOSURE 77
7.1 OVERVIEW 77
7.2 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 77
7.2.1 Air 77
7.2.2 Water 78
7.2.3 Soil 78
7.2.4 Other 78
7.3 OCCUPATIONAL EXPOSURES 79
7.4 POPULATIONS AT HIGH RISK 79
8. ANALYTICAL METHODS 81
8.1 ENVIRONMENTAL MEDIA 81
8.2 BIOMEDICAL SAMPLES 81
9. REGULATORY AND ADVISORY STATUS 85
9.1 INTERNATIONAL 85
9.2 NATIONAL 85
9.2.1 Regulations 85
9.2.2 Advisory Guidance 86
9.2.2.1 Air 86
9.2.3 Data Analysis 87
9.2.3.1 Reference dose 87
9.2.3.2 Carcinogenic potency 87
9.3 STATE 87
10. REFERENCES 89
11. GLOSSARY Ill
APPENDIX: PEER REVIEW 115
vil
-------
LIST OF FIGURES
1.1 Health effects from breathing chloroform 4
1.2 Health effects from ingesting chloroform 5
1.3 Health effects from skin contact with chloroform 6
2.1 Effects of chloroform-- inhalation exposure 11
2.2 Effects of chloroform--oral exposure 12
2.3 Effects of chloroform--dermal exposure 13
2.4 Levels of significant exposure for chloroform-- inhalation .... 14
2.5 Levels of significant exposure for chloroform--oral IS
2.6 Levels of significant exposure for chloroform--dermal 16
2.7 Availability of information on health effects of chloroform
(human data) 28
2.8 Availability of information on health effects of chloroform
(animal data) 29
4.1 Metabolic pathways of chloroform biotransformation 42
ix
-------
LIST OF TABLES
2.1 Relationship of chloroform concentration in inspired air
and blood to anesthesia 24
3.1 Chemical identity of chloroform 34
3.2 Physical and chemical properties of chloroform 35
4.1 Genotoxicity of chloroform in vitro 59
4 2 Genotoxicity of chloroform in vivo 60
4.3 Oral carcinogenicity studies of chloroform 65
6.1 Sources of chloroform released to the environment 74
8.1 Analytical methods for chloroform 82
XI
-------
1. PUBLIC HEALTH STATEMENT
1.1 WHAT IS CHLOROFORM?
Chloroform is a colorless or water-white liquid with a pleasant
nonirritating odor. Although it is both a man-made and naturally
occurring compound, human activity is responsible for most of the
chloroform found in the environment. Host of the chloroform manufactured
in the United States (93%) is used to make fluorocarbon-22.
Fluorocarbon-22 is used to make fluoropolymers and as a cooling fluid in
air conditioners. The remaining 7% of the chloroform produced in the
United States is either exported to other countries, used in the
manufacture of pesticides or dyes, or used in various products including
fire-extinguishers, dry cleaning spot removers, and various solvents.
1.2 HOW MIGHT I BE EXPOSED TO CHLOROFORM?
The general population may be exposed to chloroform by breathing
air and ingesting drinking water, beverages, and foods contaminated with
chloroform. In addition, skin contact may occur during the use of
various consumer products containing this compound or from exposure to
chlorinated waters (i.e., bath water, swimming pool water).
The primary sources of chloroform release to the environment are
pulp and paper mills, pharmaceutical manufacturing plants, chemical
manufacturing plants, chlorinated wastewater from sewage treatment
plants, and chlorinated drinking water (water is chlorinated for
disinfection purposes). Minor sources of chloroform release include, but
are not limited to, automobile exhaust gas, use of chloroform as a
pesticide, burning of tobacco products treated with chlorinated
pesticides, evaporation during shipping and transport of chloroform,
decomposition of trichloroethylene (a man-made product used primarily as
a solvent), evaporation from chlorinated tap water during showering,
evaporation from chlorinated swimming pool water, biological production
of chloroform from marine algae, reaction of chlorinated pollutants with
decayed vegetation, and burning of plastics. Most of the chloroform
released to the environment eventually enters the atmosphere, while much
smaller amounts enter groundwater as the result of filtration through
soil. Once in the atmosphere, chloroform may be transported long
distances before it finally decomposes. Chloroform present in soil may
come from improper land disposal of waste material containing chloroform
or other chlorine-containing compounds that are broken down to form
chloroform.
People who work in businesses or industries where chloroform is
found may be exposed to greater amounts of this compound than are
members of the general population. Chloroform is found in a wide variety
of occupational settings as a result of its direct use in manufacturing
-------
2 Section 1
processes, its use as a solvent for many different materials, and its
formation during various chlorination processes.
Occupational settings in which chloroform exposure may occur
include:
• Chloroform manufacturing plants
• Fluorocarbon-22 manufacturing plants
• Ethylene dichloride manufacturing plants
• Internal combustion engine industries
• Pesticide manufacturing plants
• Pulp and paper mills
• Food processing industries
• Paint stores (as a result of using chloroform-containing solvents
for lacquers, gums, greases, waxes, adhesives, oils, and rubber)
1.3 HOW DOES CHLOROFORM GET INTO MY BODY?
Chloroform can enter the body by breathing air, eating food, or
drinking water that contains chloroform. Chloroform readily penetrates
the skin; therefore, chloroform may also enter the body by bathing or
showering in water containing chloroform. Foods such as seafood, dairy
products, meat, vegetables, bread, and beverages may contain small but
measurable amounts of chloroform. Drinking-water supplies containing
organic contaminants may contain chloroform as a by-product of
chlorination of the water supply for disinfection purposes.
1.4 HOV CAR CHLOROFORM AFFECT MY HEALTH?
Chloroform affects the central nervous system, liver, and kidneys.
It was used as a surgical anesthetic for many years before its harmful
effects on the liver and kidneys were recognized. Short-term exposure to
high concentrations of chloroform in the air causes tiredness,
dizziness, and headache. Longer-term exposure to high levels of
chloroform in the air, or in food and drinking water, can affect liver
and kidney function. Toxic effects may include jaundice and burning
urination. High doses of chloroform have also been found to cause liver
and kidney cancer in experimental animals. The risks of cancer, if any,
from low-level exposures to chloroform in drinking water as a result of
chlorination, however, are far outweighed by the benefits of
chlorination in terms of greatly decreased incidence of waterborne
diseases, themselves a potential public health threat and a previous
major contributor to sickness and death.
-------
Public Health Statement 3
1.5 IS THERE A MEDICAL TEST TO DETERMINE IF I HAVE BEEN
EXPOSED TO CHLOROFORM?
Although chloroform can be detected in blood, urine, and body
tissues, the methods are not very reliable because chloroform is rapidly
eliminated from the body. In addition, the presence of chloroform in
these tissues may result from the biological breakdown of other
chlorine-containing compounds; therefore, an elevated level of
chloroform in the tissues may reflect exposure to the other compounds
rather than to chloroform itself. Measurements of blood for levels of
liver enzymes can indicate if the liver has been damaged but do not
specifically indicate if chloroform exposure occurred.
1.6 WHAT LEVELS OF EXPOSURE HAVE RESULTED IN HARMFUL HEALTH EFFECTS?
The graphs on the following pages show the relationship between
exposure to chloroform and known health effects. In health effects from
breathing chloroform (Fig. 1.1), exposure is measured in parts of
chloroform per million parts of air (ppm). In Figs. 1.2 and 1.3, the
same relationship is represented for the known health effects from
ingesting chloroform or skin contact with chloroform. Exposures are
measured in milligrams of chloroform per kilogram of body weight per
day.
The first column on these graphs, labeled "short term," refers to
known health effects in laboratory animals and humans from exposure to
chloroform for 2 weeks or less. The column labeled "long term" refers to
chloroform exposures of longer than 2 weeks. In all graphs, effects in
animals are shown on the left side and effects in humans on the right
side. The levels marked on the graphs as anticipated to be associated
with minimal risk for humans are based on information from animal
studies that are currently available; therefore, some uncertainty still
exists. From available data in animals, the Environmental Protection
Agency (EPA) has estimated that exposure to 1 microgram of chloroform
per cubic meter of air for a lifetime would result in 0.23 additional
cases of cancer in a population of 10,000 people and 230 additional
cases of cancer in a population of 10,000,000 people. This is the same
as saying that exposure to 1 part of chloroform in a billion parts of
air (1 ppb) for a lifetime would result in 11 additional cases of cancer
in a population of 10,000 people and 11,000 cases of cancer in a
population of 10,000,000 people. Exposure to drinking water containing 1
milligram of chloroform per liter of water for a lifetime would result
in 1.7 additional cases of cancer in a population of 10,000 people and
1700 additional cases of cancer in a population of 10,000,000 people. It
should be noted that these risk values are plausible upper-limit
estimates. Actual risk levels are unlikely to be higher and may be
lower.
-------
Section 1
SHORT-TERM EXPOSURE
(LESS THAN OR EQUAL TO 14 DAYS)
LONG-TERM EXPOSURE
(GREATER THAN 14 DAYS)
EFFECTS
IN
ANIMALS
CONC. IN
AIR
(ppm)
EFFECTS
IN
HUMANS
EFFECTS
IN
ANIMALS
CONC. IN
AIR
(ppm)
EFFECTS
IN
HUMANS
20,000
ANESTHESIA
1500
DEATH•
1000
DIZZINESS.
VERTIGO
LIVER DAMAGE-
500
LIVER DAMAGE.
EFFECTS ON •
THE UNBORN
ODOR
'THRESHOLD
100
90
8)
70
60
50
.20
10
TIREDNESS.
> DEPRESSION.
BURNING URINATION
0.25.
0
LIVER EFFECTS
MINIMAL RISK FOR
•EFFECTS OTHER THAN
CANCER
Fig. 1.1. Heahfc effects tnm breatfctaf cfclonfi
-------
Public Health Statement 5
SHORT-TERM EXPOSURE
(LESS THAN OR EQUAL TO 14 DAYS)
LONG-TERM EXPOSURE
(GREATER THAN 14 DAYS)
EFFECTS EFFECTS EFFECTS EFFECTS
IN DOSE IN IN DOSE IN
ANIMALS (mg/kg/day) HUMANS ANIMALS (mg/kg/day) HUMANS
2(
nCATU
1(
9
8
7
ft
(
LIVER AND
KIDNEY < <
DAMAGE
2
^2
1
(
)0 DECREASED I 100
I LONGEVITY—^ I
10 ^90
3 8
3 ri
3 6
0 5
LIVER AND
3 KIDNEY < 4
DAMAGE
0 3
L
o va
) i
•
1 0.
(
)
3
0
)
3
3
LIVER AND KIDNEY
DAMAGE
)
)
)3 MINIMAL RISK FOR
EFFECTS OTHER THAN
CANCER
)
f
-------
Section 1
SHORT-TERM EXPOSURE
(LESS THAN OR EQUAL TO 14 DAYS)
LONG-TERM EXPOSURE
(GREATER THAN 14 DAYS)
EFFECTS
IN
ANIMALS
KIDNEY —
EFFECTS
EFFECTS EFFECTS EFFECTS
DOSE IN IN IN
(mg/kg/day) HUMANS ANIMALS HUMANS
11
9
8
7
e
5
4
3
2
1
r
)0 QUANTITATIVE QUANTITATIVE
DATA WERE NOT DATA WERE NOT
AVAILABLE AVAILABLE
9
9
9
0
3
9
9
3
9
I
QUANTITATIVE
DATA WERE NOT
AVAILABLE
Fig. 1J. Health effects tnm skim caatoct with cUorofonk
-------
Public Health Seacement 7
1.7 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT
MADE TO PROTECT HUMAN HEALTH?
The government has made recommendations to limit exposure of
workers to chloroform in the workplace and exposure of the general
public to chloroform in drinking water. The National Institute for
Occupational Safety and Health (NIOSH) recommended an occupational
exposure limit of 2 parts chloroform per million parts of air averaged
over an 8-hour workday, 40-hour workweek. The Occupational Safety and
Health Administration (OSHA) has a legally enforcible ceiling limit of
50 parts-per-million chloroform in the work atmosphere, which is not co
be exceeded at any time.
EPA has promulgated a drinking water maximum contaminant level for
total trihalomethanes (compounds similar to and including chloroform) of
100 parts per billion parts of water as a technically and economically
feasible level for municipal water supplies serving 10,000 or more
individuals.
-------
2. HEALTH EFFECTS SUMMARY
2.1 INTRODUCTION
This section summarizes and graphs data on the health effects
concerning exposure to chloroform. The purpose of this section is to
present levels of significant exposure for chloroform based on key
toxicological studies, epidemiological investigations, and environmental
exposure data. The information presented in this section is critically
evaluated and discussed in Sect. 4, Toxicological Data, and Sect. 7,
Potential for Human Exposure.
This Health Effects Summary section comprises two major parts.
Levels of Significant Exposure (Sect. 2.2) presents brief narratives and
graphics for key studies in a manner that provides public health
officials, physicians, and other interested individuals and groups with
(1) an overall perspective of the toxicology of chloroform and (2) a
summarized depiction of significant exposure levels associated with
various adverse health effects. This section also includes information
on the levels of chloroform that have been monitored in human fluids and
tissues and information about levels of chloroform found in
environmental media and their association with human exposures.
The significance of the exposure levels shown on the graphs may
differ depending on the user's perspective. For example, physicians
concerned with the interpretation of overt clinical findings in exposed
persons or with the identification of persons with the potential to
develop such disease may be interested in levels of exposure associated
with frank effects (Frank Effect Level, FEL). Public health officials
and project managers concerned with response actions at Superfund sites
may want information on levels of exposure associated with more subtle
effects in humans or animals (Lowest-Observed-Adverse-Effect Level,
LOAEL) or exposure levels below which no adverse effects (No-Observed-
Adverse-Effect Level. NOAEL) have been observed. Estimates of levels
posing minimal risk to humans (Minimal Risk Levels) are of interest to
health professionals and citizens alike.
Adequacy of Database (Sect. 2.3) highlights the availability of key
studies on exposure to chloroform in the scientific literature and
displays these data in three-dimensional graphs consistent with the
format in Sect. 2.2. The purpose of this section is to suggest where
there might be insufficient information to establish levels of
significant human exposure. These areas will be considered by the Agency
for Toxic Substances and Disease Registry (ATSDR), EPA, and the National
Toxicology Program (NTP) of the U.S. Public Health Service in order to
develop a research agenda for chloroform.
-------
10 Section 2
2.2 LEVELS OF SIGNIFICANT EXPOSURE
To help public health professionals address the needs of persons
living or working near hazardous waste sites, the toxicology data
summarized in this section are organized first by route of exposure--
inhalation, ingestion, and dermal--and then by toxicological end points
that are categorized into six general areas--lethality, systemic/target
organ toxicity, developmental toxicity, reproductive toxicity, genetic
toxicity, and carcinogenicity. The data are discussed in terms of three
exposure periods--acute, intermediate, and chronic.
Two kinds of graphs are used to depict the data. The first type is
a "thermometer" graph. It provides a graphical summary of the human and
animal toxicological end points (and levels of exposure) for each
exposure route for which data are available. The ordering of effects
does not reflect the exposure duration or species of animal tested. The
second kind of graph shows Levels of Significant Exposure (LSE) for each
route and exposure duration. The points on the graph showing NOAELs and
LOAELs reflect the actual doses (levels of exposure) used in the key
studies. No adjustments for exposure duration or intermittent exposure
protocol were made.
Adjustments reflecting the uncertainty of extrapolating animal data
to man, intraspecies variations, and differences between experimental vs
actual human exposure conditions were considered when estimates of
levels posing minimal risk to human health were made for noncancer end
points. These minimal risk levels were derived for the most sensitive
noncancer end point for each exposure duration by applying uncertainty
factors. These levels are shown on the graphs as a broken line starting
from the actual dose (level of exposure) and ending with a concave-
curved line at its terminus. Although methods have been established to
derive these minimal risk levels (Barnes et al. 1987), shortcomings
exist in the techniques that reduce the confidence in the projected
estimates. Also shown on the graphs under the cancer end point are low-
level risks (lO'4 to 10'7) reported by EPA. In addition, the actual dose
(level of exposure) associated with the tumor incidence is plotted.
2.2.1 Key Studies and Graphical Presentations
Dose-response-duration data for the toxicity and carinogenicity of
chloroform are displayed in two types of graphs. These data are derived
from the key studies described in the following sections. The
"thermometer" graphs In Figs. 2.1 through 2.3 plot exposure levels vs
NOAELs and LOAELs for various effects and durations of inhalation and
oral exposures, respectively. The graphs of levels of significant
exposure in Figs. 2.4 through 2.6 plot end point-specific NOAELs and
LOAELs, and minimal levels of risk for acute (314 days), intermediate
(15-364 days), and chronic (fe365 days) durations for inhalation, oral,
and dermal exposures, respectively.
2.2.1.1 Inhalation
Lethality and decreased longevity. Data regarding inhalation
exposure levels that produce death in humans were not available. An
inhalation LCso of 10,000 ppm for 4 h for rats was reported by Lundberg
et al. (1986). Deringer et al. (1953) found that an inhalation exposure
-------
Health Effects Summary 11
ANIMALS
100000 I-
10000
1000
100
10
HUMANS
(PP«n>
100000 r
• RAT LCW. 4 h. CONTINUOUS
• CAT. CNS EFFECTS 5 MN. CONTINUOUS
• MOUSE. CNS EFFECTS. 30 MIN, CONTINUOUS
• MOUSE. CNS EFFECTS. 1 h CONTINUOUS
O MOUSE CNS EFFECTS 2 h. CONTINUOUS
- • MOUSE. DEATH KIDNEY TOXICITY 1-3 h. CONTINUOUS
• MOUSE REPRODUCTIVE EFFECTS 5 DAYS INTERMITTENT
• RAT LIVER TOXICITY 4 h CONTINUOUS
• MOUSE LIVER TOXICITY 4 h CONTINUOUS
RAT AND MOUSE. FRANK DEVELOPMENTAL TOXICITY
10 DAYS INTERMITTENT
O RAT DEVELOPMENTAL TOXICITY 10 DAYS INTERMITTENT
• RAT RABBIT GUINEA PIG LIVER AND KIDNEY EFFECTS
6 MONTHS INTERMITTENT
10000
1000
100
10
ANESTHESIA ACUTE
A ANESTHESIA ACUTE
A DIZZINESS TIREDNESS
HEADACHES
VERTIGO 3 MIN
A DIZZINESS
VERTIGO 30 MIN
SLIGHT CNS EFFECTS
OCCUPATIONAL
LONG-TERM
A LIVER TOXICITY
OCCUPATIONAL
LONG-TERM
i LOAEL FOR ANIMALS O NOAEL FOR ANIMALS
, LOAEL FOR HUMANS A NOAEL FOR HUMANS
Fif.2.1. Effects of chloroform—iahaladoo exponre.
-------
12 Section 2
ANIMALS
(mg kg/day)
1000
HUMANS
(mg/kg/day)
1000 r~
100
• MOUSE DECREASED LONGEVITY 78 WEEKS
• RAT LD». SINGLE DOSE
• RAT REPRODUCTIVE TOXICITY 13 WEEKS
• RAT. DECREASED SURVIVAL 90 DAYS
O MOUSE DECREASED LONGEVITY 78 WEEKS
/O RAT REPRODUCTIVE TOXICITY 13 WEEKS
!• MOUSE. DECREASED SURVIVAL. 6 WEEKS RAT UVER AND KIDNEY
TOXICITY 13 WEEKS
• MOUSE L0» SINGLE DOSE
• RAT DECREASED LONGEVITY LIVER TOXICITY 78 WEEKS
f O RAT DECREASED SURVIVAL. 90 DAYS
< O MOUSE DECREASED SURVIVAL 6 WEEKS RAT LIVER TOXICITY 80 WEEKS
I • MOUSE KIDNEY TOXICITY 80 WEEKS
I • RAT LIVER TOXICITY 10 DAYS MOUSE LIVER TOXICITY 90 DAYS
IO RABBIT DEVELOPMENTAL TOXICITY 12 DAYS
• MOUSE. LIVER TOXICITY 90 DAYS
O MOUSE. UVER TOXICITY 90 DAYS
• MOUSE. KIDNEY TOXICITY 14 DAYS
[• DOG LIVER TOXICITY. 18 WEEKS
\ • MOUSE UVER TOXICITY SINGLE DOSE
lO RAT LIVER AND KIDNEY TOXICITY 13 WEEKS
O RAT UVER TOXICITY 10 DAYS
O MOUSE UVER AND KIDNEY TOXICITY SINGLE DOSE
O MOUSE MONEY TOXICITY 80 WEEKS
• DOG LIVER TOXICITY 7 S YEARS
100 —
10
10 «—
0 1
A DEATH SINGLE
DOSE
LIVER AND
KIDNEY
TOXICITY
1-5 YEARS
• LOAEL FOR ANIMALS
O NOAEL FOR ANIMALS
LOAEL FOR HUMANS
NOAEL FOR HUMANS
Fif.24. Effects of cUorofofB—onl
-------
Health Effects Summary 13
ANIMALS
(mg/kg/day)
100 i- • RABBIT. KIDNEY TOXICITY. 24 h
HUMANS
10
QUANTITATIVE DATA
WERE NOT AVAILABLE
LQAEL
Fig. 2J. Effects of chloroform—dermal exposure.
-------
14 Section 2
ACUTE INTERMEDIATE CHRONIC
(S14 DAYS) (15-364 DAYS) (2365 DAYS)
DEVELOP- TARGET REPRO- TARGET TARGET
LETHALITY MENTAL ORGAN DUCTIVE ORGAN ORGAN CANCER
(ppm)
10.000
r«r
1000
100
10
0 1
001
0001
00001
000001
0 00000T
-•m
r RAT
m MOUSE
g GUINEA PIG
h RABBIT
• m
• r (LIVER)
r • m • m (LIVER)
r. h. g (LIVER AND KIDNEY)
A (LIVER)
I
LOAEL AND NOAEL
IN THE SAME
SPECIES
! MINIMAL RISK
« LEVEL FOR
"^EFFECTS OTHER
THAN CANCER
,-4
10
ID'5 ESTIMATED
UPPER-BOUND
HUMAN
CANCER
10-6 RISK LEVELS
10
,-7
• LOAEL FOR ANIMALS
O NOAEL FOR ANIMALS
A LOAEL FOR HUMANS
Fig. 2.4. Leveb of «ig»ifl>«ii« exporare for chloroform—iakdatioa.
-------
Health Effects Summary 15
ACUTE
(S14 DAYS)
DEVELOP
LETHALITY MENTAL
TARGET
ORGAN
INTERMEDIATE
(15-364 DAYS)
HEPRO-
LETHALITY DUCTIVE
TARGET
ORGAN
CHRONIC
<> 36S DAYS)
DECREASED TARGET
LONGEVITY ORGAN CANCER
(mg kg/day)
1000
100
10
0 1
001
0001
00001
000001
Oh
r (LIVER)
m(UVER)
•
» T
r RAT
m MOUSE
h RABBIT
d 000
I r ,
, MINIMAL RISK LEVEL
' FOR EFFECTS OTHER
^ THAN CANCER
I
LOAEL AND NOAEL IN
SAME SPECIES
m(UVER)f
0
f
T
1
0
' (LIVER
K.DNEY)
«d (LIVER)
• m
0 r (LIVER)
.
6
• LOAEL FOR ANIMALS
O NOAEL FOR ANIMALS
A LOAEL FOR HUMANS
A NOAEL FOR HUMANS
« m
.,
(KIDNEY) • m
1 • '
° f d (LIVER)
(LIVER) A
10" -
10-' -
ESTIMATED
UPPER-BOUND
HUMAN
CANCER
RISK LEVELS
Fig. L5. Levcb of significant exposure for chlorofonn—oraL
-------
16 Section 2
(mg/kg/day)
100 r-
10
ACUTE
(514 DAYS)
TARGET ORGAN
• h (KIDNEY)
INTERMEDIATE
(15-364 DAYS)
QUANTITATIVE DATA
WERE NOT
AVAILABLE
CHRONIC
(* 365 DAYS)
QUANTITATIVE DATA
WERE NOT
AVAILABLE
h RABBIT
• LOAEL FOR ANIMALS
Fig. 2.6. Lefcbofripfflcut
for
-------
Health Effects Summary 17
of 1025 ppm for 1 to 3 h was fatal to mice. These FELs are displayed as
LOAELs under acute exposure in Figs. 2.1 and 2.4. No data were available
regarding decreased longevity of animals due to longer-term inhalation
exposure to chloroform.
Systemic/target organ toxicity. Target organs for chloroform
toxicity are the liver, kidney, and central nervous system (CNS). CNS
toxicity is usually observed at very high exposures. In humans, levels
of 20,000 to 40,000 ppm were used to produce anesthesia (NIOSH 1974);
levels <1SOO ppm are insufficient to produce anesthesia (Goodman and
Oilman 1980); dizziness and vertigo occurred at 920 ppm for 3-min
(Lehman and Hasegawa 1910, Lehman and Schmidt-Kehl 1936), and no
symptoms were reported by subjects exposed to 390 ppm for 30 min (EPA
1985a). For longer-term exposures, occupational exposure to >77 ppm
resulted in symptoms of tiredness and depression, whereas levels of 22
co 71 ppm resulted in less severe manifestations of these symptoms
(Challen et al. 1958). CNS effects in animals include disturbed
equilibrium in cats at 7200 ppm for 5 min. deep narcosis in mice at 4000
ppm for 30 min, slight narcosis in mice at 3100 ppm for 1 h, and no
obvious effects at 2500 ppm for 2 h (EPA 1985a). These levels are
plotted in Fig. 2.1, but not in Fig. 2.4 because effects on the liver
and kidney occur at lower exposure levels, and only the most sensitive
target organ effect for each species is plotted in Fig. 2.4.
The liver and kidney are the most sensitive targets for systemic
toxicity. Phoon et al. (1983) reported cases of toxic jaundice among
factory workers at occupational exposures of 14.4 to >400 ppm. High
incidences of toxic hepatitis and liver enlargement were found in
workers exposed to chloroform at levels of 2 to 205 ppm for 1 to 4 years
(Bomski et al. 1967). The lower end of this range (2 ppm) is plotted in
Figs. 2.1 and 2.4. However, because the exact exposure level associated
with liver effects in humans exposed by inhalation cannot be determined.
no minimal risk level was derived from these data.
Liver (and kidney) effects have been observed in rats and mice
after acute inhalation exposures, and in rats, rabbits, and guinea pigs
after intermediate exposure. None of the studies define NOAELs. In a
study by Kyiin et al. (1963), fatty infiltration occurred in mice at 100
ppm for 4 h (LOAEL), and dose-related liver necrosis occurred at 200,
400, and 800 ppm (FELs). Thus, 100 ppm is a short-term inhalation LOAEL
for liver effects in mice and is plotted in Figs. 2.1 and 2.4 for acute
exposure. A 4-h TC50 of 120 ppm for liver damage was reported for rats
and is represented as an acute LOAEL for liver effects in rats in Figs.
2.1 and 2.4. For intermediate exposure, 25 ppm, 7 h/day, 5 days/week for
6 months is a LOAEL for liver effects (lobular granular degeneration and
focal necrosis) and kidney effects (cloudy swelling) in rats, rabbits,
and guinea pigs in a study by Torkelson et al. (1976). Higher levels (50
and 85 ppm) resulted in more severe effects in rats. The 25-ppm level is
plotted in Figs. 2.1 and 2.4 as a LOAEL for target organ effects for
intermediate exposure and is the basis for the minimal risk level for
intermediate inhalation exposure. Since most studies in animals were
conducted to provide information relevant to the clinical use of
chloroform as an anesthetic, no chronic inhalation study data were
available.
-------
18 Section 2
Developmental toxicity. No data regarding human developmental
effects of chloroform were available. A study by Schwetz et al. (1974)
defines a NOAEL and FELs for developmental effects in rats exposed by
inhalation. Exposure of pregnant rats to 30 ppm (NOAEL) 7 h/day on days
6 to 15 of gestation resulted in no effects on the offspring, whereas
100 ppm (FEL) caused increased incidences of missing ribs, imperforate
anus, subcutaneous edema, and delayed ossification of sternebrae, and
300 ppm (FEL) caused abnormalities of the skull and sternum, decreased
numbers of live fetuses/litter, and increased resorptions. The NOAEL and
lower FEL for rats are plotted in Figs. 2.1 and 2.4 under acute
exposure. The NOAEL is the basis for the minimal risk level for acute
inhalation exposure. Exposure of pregnant mice on various days of
gestation to 100 ppm for 7 h/day resulted in frank effects, such as
increased numbers of resorptions and increased incidences of cleft
palate (Murray et al. 1979). This level in mice is also indicated in
Figs. 2.1 and 2.4.
Reproductive tozicity. The only data regarding reproductive
effects of inhalation exposure to chloroform are that exposure of male
mice to 400 or 800 ppm, 4 h/day for 5 days caused significant increases
in the percentage of abnormal sperm (Land et al. 1981) (see Figs. 2.1
and 2.4).
Genotoxicity. Mixed results were obtained in sister chromatid
exchange (SCE) assays in cultured human lymphocytes (Sect. 4.3.5 on
genotoxicity in toxicological data section). Studies on the in vitro
genotoxicity of chloroform reported negative results in bacteria, mixed
results in yeasts, and negative results for gene mutations and
chromosome aberrations in mammalian cells. In vivo gene mutation tests
in Drosophila and DNA damage in rats and mice were negative, whereas
tests for chromosome aberrations and sperm abnormalities were mixed.
Carcinogenicity. Data regarding the carcinogenicity of inhaled
chloroform in humans and animals were not available. Studies in animals
indicate that chloroform is carcinogenic by the oral route. NCI (1976)
found dose-related increased incidences of hepatocellular carcinoma in
male and female mice treated by gavage at time-weighted average (TWA)
doses of &138 mg/kg/day 5 days/week for 78 weeks, and a dose-related
increased incidence of kidney epithelial tumors in male rats similarly
treated by gavage at 90 and 180 mg/kg/day. Roe et al. (1979) found an
increased incidence of kidney epithelial tumors in male mice given 60
mg/kg/day 6 days/week for 78 weeks. Dose-related increased Incidences of
renal tubular cell adenomas and/or carcinomas were found in male rats
treated with chloroform in the drinking water at levels equivalent to
dosages 238 mg/kg/day for 104 weeks (Jorgenaon et al. 1985).
The EPA (1985a) considered these five data sets in determining the
q * for chloroform. The five data sets were (1) liver tumors in female
mice (NCI 1976), (2) liver tumors in male mice (NCI 1976). (3) kidney
tumors in male rats (NCI 1976), (4) kidney tumors in male mice (Roe et
al. 1979), and 5) kidney tumors in male rats (Jorgenson et al. 1985).
EPA (1985a) used available pharmacokinetic data to calculate an
effective dose for these studies, assuming that the amount metabolized
to reactive metabolites is the gavage dose minus the amount excreted
unchanged. For mice given 60 mg/kg, as in the Roe et al. (1979) study,
-------
Health Effects Summary 19
the correction was 6%. For rats at the sane dosage, it was 20%. In the
NCI (1976) study in which rats and mice received doses of -200 to 500
mg/kg/day, a 20% correction was considered conservative and would
probably overestimate the amount metabolized from these doses. EPA
(198Sa) used these correction factors to reduce the administered dose by
the unmetabolized portion (6% in mice and 20% in rats when given as a
bolus by gavage in corn oil, 0% when administered in drinking water).
Doses were also corrected for differences between animal and human
pharmacokinetics by using a surface area correction. Using these
corrected doses, maximum likelihood estimates of the parameters of the
multistage model were calculated for each of the five data sets. EPA
(198Sa) chose the mouse liver tumor data from the NCI (1976) study as
the basis of the potency factor for inhalation exposure to chloroform.
The NCI (1976) study is considered to be appropriate for use in the
inhalation risk estimate because there were no inhalation cancer
bioassays and no pharmacokinetic data to contraindicate the use of
gavage data (EPA 1987b). The geometric mean of the estimates for male
and female mice in the NCI (1976) study, 8.1 x 10'2 (mg/kg/day) *1. was
recommended as the inhalation q.* for chloroform. EPA (198Sa) combined
the estimates for both data sets because the data for males included
observations at a lower dose, which appeared to be consistent with the
female data. EPA (1985a) noted that the recommended q * was similar to
the geometric mean calculated from all five estimates and was also
similar to the estimate calculated if data for both sexes of B6C3F1 mice
in the NCI (1976) study were pooled. Expressed in terms of concentration
in air, the qi* is equal to 2.3 x 10'5 (jig/m3)'1 or 1.1 x 10'4 (ppb)'1.
The concentrations in air associated with individual lifetime
upperbound risks of 10'4, 10'5, 10'6, and 10'7 are 4.3 x 10'3 4.3 x
10'4, 4.3 x ID'5, and 4.3 x 10'6 mg/m3 (8.8 x 10'4, 8.8 x 10'5, 8.8 x
10'6, and 8.8 x 10*7 ppm), respectively, assuming that a 70-kg human
breathes 20 m3 air/day. The 10'4 to 10'7 levels are indicated in Fig.
2.1.
2.2.1.2 Oral
Lethality and decreased longevity. A fatal oral dose of chloroform
may be as little as 10 mL (14.8 g or 211 mg/kg for a 70-kg human)
(Schroeder 1965). This dose is plotted in Figs. 2.2 and 2.5 for acute
lethality. Longer-term exposure levels causing death in humans were not
available.
A wide range of oral U>SO values have been reported for rats and
mice. The lowest were 444 mg/kg in 14-day-old rats (Kimura et al. 1971)
and 118 mg/kg in an especially sensitive strain of mice (Hill 1978).
These levels are plotted in Figs. 2.2 and 2.5 for acute lethality.
Increased mortality was reported in oral studies of intermediate
duration. In mice, gavage doses of £150 mg/kg/day, but not £60
mg/kg/day, 6 days/week for 6 weeks resulted in Increased rates of
mortality (Roe et al. 1979). Thus, 150 mg/kg/day is the lowest FEL and
60 mg/kg/day is the NOAEL for increased mortality for intermediate oral
exposure in mice. In rats, 90-day exposure to 2500 ppm in drinking water
resulted in increased mortality, whereas exposure to 3500 ppm did not
(Chu et al. 1982a). Assuming that a 0.35-kg rat consumes 0.049 L of
-------
20 Section 2
water per day (EPA 1985b), the 2500-ppm PEL and 500-ppm NOAEL for
mortality for Intermediate oral exposure in rats are equivalent to doses
of 350 and 70 mg/kg/day. respectively. Levels are plotted in Figs. 2 2
and 2.5.
For chronic exposures, effects on survival were observed in studies
in rats and mice. In rats, gavage doses of a90 mg/kg/day for 78 weeks
followed by 33 weeks observation resulted in dose-related increased
mortality, due perhaps to liver toxicity (NCI 1976). In mice, a gavage
dose of 477 mg/kg/day for 78 weeks followed by 14 to 15 weeks
observation resulted in decreased survival, whereas doses <238 mg/kg/day
did not (NCI 1976). The FELs and NOAELs for increased mortality for
chronic oral exposure for rats and mice are indicated in Figs 2 2
and 2.5.
As seen from Fig. 2.5, increased mortality in mice in chronic
studies did not appear to occur at gavage doses that resulted in death
in acute experiments. The discrepancy is explained primarily by strain
differences in sensitivity to chloroform.
Systemic/target organ tozicity. The liver and kidney are the
target organs of oral exposure to chloroform. Levels of acute oral
exposure that result in liver and kidney effects in humans were not
available. A patient who ingested 1.6 to 2.6 g/day in cough medicine for
-10 years developed hepatitis and nephrosis (Wallace 1950). The presence
of other materials in the cough medicine and the fact that the subject
also ingested moderate amounts of alcohol, a known liver toxicant, make
this study unsuitable for derivation of minimal risk levels. No evidence
of liver or kidney disease, based on liver and kidney function tests,
were found in volunteers who were exposed for 1 to 5 years to -68 mg/day
chloroform in a dentifrice or -197 mg/day in a dentifrice and mouthwash
(OeSalva et al. 1975). The authors calculated an equivalent ingestion of
0.96 mg/kg/day for a 50-kg adult exposed to the higher dosage, which was
assumed to be 25% ingested. The dosage of 0.96 mg/kg/day is a chronic
oral NOAEL for liver and kidney effects. The human NOAEL is indicated in
Figs. 2.2 and 2.5 for chronic oral target organ toxicity.
Two acute studies define a LOAEL and a NOAEL for liver effects in
mice. Jones et al. (1958) observed fatty infiltration in the livers of
mice after a single dose of 30 mg/kg (LOAEL) and centrilobular necrosis
at 133 mg/kg (FEL). Moore et al. (1982) found no toxic effects on the
livers of mice after single doses of 18 mg/kg (NOAEL) or 60 mgAg, but
increased serum glutaaic-oxaloacetic transaminase (SCOT) occurred at 199
mg/kg- The 18 mgAg dose is also a NOAEL for kidney effects in mice, but
at 60 mg/kg/day, there was increased kidney weight, tubular necrosis.
and tubular regeneration. The acute oral NOAEL (18 mgAg) and LOAEL (30
mgAg) for target organ toxicity in mice are Indicated on Figs. 2.2 and
2.5. The NOAEL is the basis for the minimal risk level for acute oral
exposure.
Condie et al. (1983) observed adverse effects on the kidneys of
mice administered gavage doses of fc37 mgAg/day chloroform in corn oil
for 14 consecutive days. The lowest dose used in this study, 37
"gAg/day, is a LOAEL for kidney effects in short-term oral exposure in
mice (Fig. 2.2). Because this is not the most sensitive target organ
-------
Health Effects Summary 21
effect observed in short-term oral exposures in mice, it is not
presented in Fig. 2.5.
A teratology study provides dose-response data for liver effects
for short-term oral exposure of rats. No effect occurred in pregnant
rats treated at 20 mg/kg/day (NOAEL) for 10 days during gestation, fatty
infiltration occurred at SO mg/kg/day (LOAEL), and hepatitis occurred at
316 mg/kg/day (PEL) (Thompson et al. 1974). The acute oral NOAEL and
LOAEL for target organ effects in rats are indicated on Figs. 2.2 and
2.5. Short-term oral exposure of rats also resulted in kidney effects,
but only at doses much higher than those that caused liver effects
(Sect. 4.3.2.1 on systemic/target organ toxicity in the toxicological
data section).
Intermediate duration studies define a NOAEL and LOAEL for liver
effects in mice for oral exposure. In a study by Hunson et al. (1982),
the lowest dose tested, 50 mg/kg/day (LOAEL) by gavage for 90 days,
resulted in increased relative and absolute liver weights accompanied by
slight histopathological changes. In a drinking water study, exposure Co
200 ppm (40 mg/kg/day) for 30 to 90 days was a NOAEL, whereas >400 ppm
(>80 mg/kg/day) resulted in mild hepatic centrilobular fatty changes
(Jorgenson and Rushbrook 1980). The intermediate oral NOAEL and LOAEL
for liver effects in mice are indicated in Figs. 2.2 and 2.5.
Intermediate exposure studies found liver and kidney effects in
rats. Gavage dosing of rats with toothpaste containing chloroform, 6
days/week for 13 weeks, resulted in fatty and necrotic liver changes ac
410 mg/kg/day and increased relative liver and kidney weight at 150
mg/kg/day (LOAEL). No effects occurred at 30 mg/kg/day (NOAEL) (Palmer
et al. 1979). The intermediate oral LOAEL and NOAEL for target organ
effects in rats are indicated in Figs. 2.2 and 2.5.
In dogs, treatment with chloroform in gelatin capsules for 18 weeks
resulted in elevated serum glutamic-pyruvic transaminase (SGPT) at 30
mg/kg/day (LOAEL), the lowest dose tested (Heywood et al. 1979). This
level is indicated in Figs. 2.2 and 2.5 for intermediate target organ
(liver) toxicity in dogs and is the basis for the minimal risk level for
intermediate oral exposure.
Chronic oral studies have reported effects on livers of rats, mice,
and dogs, and on kidneys of mice. In mice, nonneoplastic proliferative
changes and necrosis occurred in livers at £138 mg/kg/day, given by
gavage 5 days/week for 78 weeks followed by 15 weeks of observation (NCI
1976). In mice treated by gavage with chloroform 6 days/week for 80
weeks followed by 16 to 24 weeks of observation, there was an increased
incidence of moderate to severe kidney disease at 60 mg/kg/day (LOAEL),
but not at 17 mg/kg/day (NOAEL) (Roe et al. 1979). The levels are
indicated on Figs. 2.2 and 2.5.
In rats, dose-related increased incidences of liver necrosis
occurred at gavage doses £90 mg/kg/day (LOAEL), 5 days/week for 78 weeks
followed by 33 weeks of observation (NCI 1976). Minor histological
changes in the liver, without evidence of chloroform-induced
hepatotoxicity, were observed in rats treated at 60 mg/kg/day (NOAEL) in
toothpaste, 6 days/week for 80 weeks, followed by 15 weeks of
-------
22 Section 2
observation (Palmer et al. 1979). The chronic oral NOAEL and LOAEL for
target organ effects in rats are indicated in Figs. 2.2 and 2.5.
' Doses of 215 mg/kg/day given by gavage to dogs 6 days/week for 7.5
years resulted in increased levels of SGPT and other serum enzymes
indicative of liver damage, as well as increased numbers of fatty cysts
(Heywood et al. 1979). The dose of 15 mg/kg/day was the lowest dose
tested and is considered a chronic oral LOAEL for target organ effects
in dogs (see Figs. 2.2 and 2.5). The minimal risk level for chronic oral
exposure was calculated from this LOAEL.
Developmental toxicity. In a teratogenicity study, pregnant rats
were treated by gavage with chloroform at 0, 20, 50, and 126 mg/kg/day
on days 6 to 15 of gestation (Thompson et al. 1974). No effects occurred
at 20 mg/kg/day- No fetal effects occurred at 50 mg/kg/day, but maternal
toxicity was evident at this level and higher. Fetal body weights were
significantly decreased at 126 mg/kg/day. For reasons discussed in Sect.
4.3.3.2 on developmental toxicity in animals after oral exposure in the
toxicological data section, it is not appropriate to define a LOAEL and
NOAEL for developmental effects in rats from this study. Thompson et al.
(1974) also treated pregnant rabbits with 20, 35, or 50 mgAg/day on
days 6 to 18 of gestation and found no treatment-related developmental
effects. Thus, 50 mgAg/day is a NOAEL for developmental effects of
orally administered chloroform in rabbits (see Figs. 2.2 and 2.5).
Reproductive toxicity. The only data regarding reproductive
effects are that rats administered gavage doses of 410 mg/kg/day
(LOAEL), but not 150 mgAg/day (NOAEL) 6 days/week for 13 weeks, had
gonadal atrophy (Palmer et al. 1979). These levels are indicated in Fig.
2.2 and in Fig. 2.5 under intermediate exposure.
Genotoxicity. See Sect. 2.2.1.1 on genotoxicity under inhalation
exposure.
Carcinogenicity. Epidemiological studies indicate a possible
relationship between exposure to chlorinated drinking water and cancer
of the bladder, large intestine, and rectum in humans (EPA 1985a).
Chloroform is one of several volatile organic contaminants considered to
have carcinogenic potential, but it has not been identified as the sole
or primary cause of excess cancer rates associated with chlorinated
drinking water.
Studies in animals indicate that chloroform is carcinogenic. NCI
(1976) found dose-related increased incidences of hepatocellular
carcinoma in male and female mice treated with chloroform in corn oil by
gavage at a TWA dose of 138 mgAg/day and above, 5 days/week for 78
weeks, and a dose-related increased incidence of kidney epithelial
tumors in male rats treated by gavage at 90 and 180 mgAg/day 5
days/week for 78 weeks. An Increased incidence of epithelial tumors of
the kidney occurred in male mice given 60 mg/kg/day. but not 17
mgAg/day 6 days/week for 78 weeks (Roe et al. 1979). Dose-related
increased incidences of renal tubular cell adenomas and/or carcinomas
were found in male rats treated with chloroform in the drinking water at
levels equivalent to dosages 238 mgAg/day for 104 weeks (Jorgenson et
al. 1985). Doses associated with increased incidences of tumors in these
studies are indicated in Fig. 2.5.
-------
Health Effects Summary 23
EPA (1987b) chose the study by Jorgenson et al. (1985) as the basis
for the q. for oral exposure to chloroform because administration via
drinking water better approximates oral exposure of humans than does
administration in corn oil by gavage in the NCI (1976) study. Based on
the incidence of renal tumors in male Osborne-Mendel rats, the q * was
calculated to be 6.1 x 10*3 (mg/kg/day)"1. The oral doses associated
with individual lifetime upper-bound risks of
10-4 10-5f 10-6f and 10-7 are 1.6 x 10'2, 1.6 x 10'3. 1.6 x 10'4, and
1.6 x 10*5 mg/kg/day, respectively. These doses are indicated in Fig.
2.5.
2.2.1.3 Dermal
Lethality and decreased longevity. Pertinent data regarding
lethality and reduced longevity due to dermal exposure to chloroform
were not located in the available literature.
Systemic/target organ toxicity. The only information found was
that dermal applications as low as 100 mg/kg for 24 h caused
degenerative changes in kidney tubules of rabbits (Torkelson et al.
1976). This one data point is displayed graphically in Figs. 2.3 and
2.6.
Developmental tozicity. Pertinent data regarding developmental
toxicity due to dermal exposure to chloroform were not located in the
available literature.
Reproductive toxicity. Pertinent data regarding reproductive
effects of dermal exposure to chloroform were not located in the
available literature.
Genotoxicity. See Sect. 2.2.1.1 on genotoxicity under inhalation
exposure.
Carcinogenicity. Pertinent data regarding the carcinogenicity of
dermal exposure to chloroform were not located in the available
literature.
2.2.2 Biological Monitoring as a Measure of Exposure and Effects
Methods for measuring chloroform in biological fluids, tissues, and
exhaled breath are available; however, there is relatively little
quantitative information relating monitored chloroform levels in tissues
or fluids to exposure levels or toxic effects, possibly because of
chloroform's volatility and rapid elimination from tissues. In
addition, the presence of chloroform or its metabolites in biological
fluids and tissues may result from the metabolism of other chlorinated
hydrocarbons; thus, elevated tissue levels of chloroform or its
metabolites may reflect exposure to other compounds. The relationship
between chloroform concentration in inspired air and resulting blood
chloroform levels is the most well-defined measure of exposure due to
the extensive use of chloroform as a surgical anesthetic. Smith et al.
(1973) observed a mean arterial blood concentration of 9.8 mg/dL (rag %)
(range 7-16.5) among 10 patients receiving chloroform anesthesia at an
inspired air concentration of 8000 to 10,000 ppm. Similar findings were
reported by Goodman and Oilman (1980) (Table 2.1). Blood chloroform
levels were monitored by Phoon et al. (1983) in workers experiencing
-------
24 Section 2
Table 2.1. Relationship of chloroform conceotratioo
in inspired air and Mood to anesthesia
Condition
In inhaled air
(ppm)
In blood
(mg%)
Not sufficient for anesthesia < 1,500 <2
Light anesthesia (after induction) 1,500-2,000 2-10
Deep anesthesia 2,000-15,000 10-20
Respiratory failure 20,000 20-25
Source: Goodman and Oilman 1980, EPA 1985a.
-------
Health Effects Summary 25
toxic jaundice due to chloroform exposure. When workroom air
concentrations were estimated to be >400 ppm, the blood samples of 13
workers with jaundice were 0.10 to 0.20 pg/100 mL blood. In another
group of 18 workers with toxic hepatitis, blood samples revealed
chloroform in some but not all workers, and workroom air contained 14.4
to 50.4 ppm on various days. Although these data indicate a trend for
increased blood concentrations with increased exposure concentration,
the wide ranges in exposure concentrations associated with wide ranges
of blood concentrations preclude meaningful graphical display.
Studies in animals also indicate that blood chloroform level is noc
a reliable indicator of exposure in animals. Jorgenson et al. (198S)
measured blood chloroform concentrations in rats chronically exposed to
chloroform in their drinking water at concentrations of 0 to 1800 mg/L.
In addition to dose-related increased incidences of renal tubule tumors,
they observed a general dose-related trend in blood chloroform levels,
but noted that the levels were highly variable, with standard deviations
usually greater than the mean. This variability could be a result of the
very short half-life of chloroform in the blood due to pulmonary
excretion. Although monitoring blood chloroform levels appears to be the
best indication of recent exposure, the sensitivity and reliability of
this indicator are questionable.
Wallace et al. (1986a,b) and Wallace (1986) measured the levels of
chloroform in breathing zone (personal) air, outdoor air, and exhaled
breath of people residing in industrial urban and rural areas. No
significant correlation was found in chloroform levels in exhaled breach
either with personal air or outdoor levels.
The primary targets of chloroform toxicity are the CNS, liver, and
kidneys. The signs of CNS effects (e.g., dizziness, tiredness, headache)
are readily apparent. Monitoring for effects of lower-level chloroform
exposure on liver and kidneys involves functional tests. Liver effects
are commonly detected by monitoring for elevated levels of liver enzymes
in the serum. Urinalysis and measurement of blood urea nitrogen are used
to detect abnormalities in kidney function. The quantitative
relationship between blood chloroform levels and measurable effects has
not been adequately determined.
2.2.3 Environmental Levels as Indicators of Exposure and Effects
2.2.3.1 Levels found in the environment
Monitoring data for chloroform in soil were not located.
Chloroform is found in drinking water as a result of chlorination of
water containing organic matter (Cech et al. 1982, Beliar et al. 1974),
but the association between drinking water levels and biological
indicators of exposure has not been defined. Chloroform has also been
detected in various food items (McConnell et al. 1975, Entz et al. 1982,
Lovegren et al. 1979, Coleman et al. 1981, Pellizzari et al. 1982), but
data are insufficient to predict the average daily intake from dietary
sources, and an association between dietary intake and biological
indicators of exposure has not been defined.
-------
26 Section 2
2.2.3.2 Human exposure potential
The level of chloroform In chlorinated drinking waters is expected
to be highest in areas where raw water supplies have high concentrations
of humic/fulvic acids, high concentrations of algae, alkaline pH, and
relatively warm temperatures. These factors contribute to elevated
chloroform levels in finished drinking waters (EPA 1981). Treatment
techniques are available to lower the level of chloroform in finished
drinking waters; however, technology is costly, and many communities,
particularly those serving small populations, may not be able to finance
such operations.
Based on the relatively high water solubility of chloroform [8.95 x
103 mg/L at 20°C (DeShon 1979)], it is speculated that this compound
would be readily available for uptake from soil by plants. Detection of
chloroform in a wide variety of food items (HcConnell et al. 1975, Entz
et al. 1982, Lovegren et al. 1979, Coleman et al. 1981, Pellizzari et
al. 1982) appears to support this prediction. Because of the lack of
data on this topic, it is not certain what effect environmental
conditions and soil characteristics would have on chloroform uptake.
Monitoring data indicate that people living in highly populated and
industrial source-dominated areas are exposed to higher levels of
chloroform in air than people living in urban or rural nonsource-
dominated areas (Brodzinsky and Singh 1982).
2.3 ADEQUACY OF DATABASE
2.3.1 Introduct ion
Section 110 (3) of SARA directs the Administrator of ATSOR to
prepare a toxicological profile for each of the 100 most significant
hazardous substances found at facilities on the CERCLA National
Priorities List. Each profile must include the following content:
"(A) An examination, summary, and interpretation of available
toxicological information and epidemiologic evaluations on a
hazardous substance in order to ascertain the levels of
significant human exposure for the substance and the
associated acute, subacute, and chronic health effects.
(B) A determination of whether adequate information on the health
effects of each substance is available or in the process of
development to determine levels of exposure which present a
significant risk to human health of acute, subacute, and
chronic health effects.
(C) Where appropriate, an identification of toxicological testing
needed to identify the types or levels of exposure that may
present significant risk of adverse health effects in humans."
This section identifies gaps in current knowledge relevant to
developing levels of significant exposure for chloroform. Such gaps are
identified for certain health effect "end points" (lethality,
systemic/target organ toxicity, developmental toxicity, reproductive
toxicity, and cancer) reviewed in Sect. 2.2 of this profile in
developing levels of significant exposure for chloroform, and for other
-------
Health Effects Summary 27
areas such as human biological monitoring and mechanisms of toxicity.
The present section briefly summarizes the availability of existing
human and animal data, identifies data gaps, and summarizes research in
progress that may fill such gaps.
Specific research programs for obtaining data needed to develop
levels of significant exposure for chloroform will be developed by
ATSDR, NTP, and EPA in the future.
2.3.2 Health Effect End Points
2.3.2.1 Introduction and graphic summary
The availability of data for health effects in humans and animals
is depicted on bar graphs in Figs. 2.7 and 2.8, respectively.
The bars of full height indicate that there are data to meet at
least one of the following criteria:
1. For noncancer health effect end points, one or more studies are
available that meet current scientific standards and are sufficient
to define a range of toxicity from no-effect levels (NOAELs) to
levels that cause effects (LOAELs or FELs).
2. For human carcinogenicity, a substance is classified as either a
"known human carcinogen" or "probable human carcinogen" by both EPA
and the International Agency for Research on Cancer (IARC)
(qualitative), and the data are sufficient to derive a cancer
potency factor (quantitative).
3. For animal carcinogenicity, a substance causes a statistically
significant number of tumors in at least one species, and the data
are sufficient to derive a cancer potency factor.
4. There are studies which show that the chemical does not cause this
health effect via this exposure route.
Bars of half height indicate that "some" information for the end
point exists but does not meet any of these criteria.
The absence of a column indicates that no information exists for
that end point and route.
2.3.2.2 Descriptions of highlight* of graphs
Information regarding effects of acute inhalation exposure of
humans to chloroform has come from the former use of chloroform as an
anesthetic. As these exposure levels were high and resulted mainly in
CNS effects, the data do not adequately characterize other possible
systemic effects of inhalation exposure to chloroform in humans. Long-
term occupational exposure to chloroform may result in slight CNS and
liver effects, but exposure levels are not well characterized. Thus, the
bars for acute and chronic systemic toxicity due to inhalation exposure
indicate that there are some data (see Fig. 2.7). No data were available
regarding the lethality, developmental and reproductive toxicity, and
carcinogenicity of inhaled chloroform in humans. For oral exposure, some
-------
HUMAN DATA
Ni
00
•
n
>~.
§
K)
V SUFFICIENT
^INFORMATION*
SOME
INFORMATION
NO
INFORMATION
INHALATION
DERMAL
LETHALITY
ACUTE
INTERMEDIATE CHRONIC DEVELOPMENTAL REPRODUCTIVE CAHCINOOENICITY
/ TOXICITV TOXICITY
SYSTEMIC TOXICITV
'Sufficient information exists to meet at least one of the criteria for cancer or noncancer end points.
Fig. 2.7. Availability of information on health effects of chloroform (human data).
-------
ANIMAL DATA
v SUFFIOEHT
^INFORMATION*
J
SOME
'INFORMATION
NO
INFORMATION
OHAL
INHALATION
DERMAL
LETHALITY ACUTE INTERMEDIATE CHROMIC DEVELOPMENTAL REPRODUCTIVE CARCINOOENICITY
/ / TOXICITV TOXICITY
SYSTEMIC TOXICITY
'Sufficient information exists to meet at least one of the criteria for cancer or noncancer end points.
fig. 2.8. Availability of information on health effects of chloroform (animal data).
ro
\O
-------
30 Section 2
data exist for acute lethality in humans. The only data regarding
systemic toxicity due to chronic oral exposure are that daily use of
dentifrices and mouthwashes containing chloroform did not result in
evidence of liver or kidney disease. Chlorination of drinking water may
be associated with an increased risk of cancer, but chloroform is one of
several contaminants in chlorinated water and has not been identified as
the only or primary cause of excess cancer rates. Thus, bars for
lethality, chronic systemic toxicity, and carcinogenieity of oral
exposure to chloroform indicate some data (see Fig. 2.7). No data were
available for systemic toxicity of acute or intermediate oral exposure
or for developmental and reproductive toxicity of oral exposure to
chloroform in humans. No data were available for dermal exposure of
humans to chloroform.
In animals, LOSQs for rats and mice were available for oral
exposure. Doses that caused death and that did not cause death were
available for rats and mice for intermediate oral exposure and for mice
for chronic oral exposure, and a LOAEL for decreased longevity for
chronic oral exposure of rats was defined. Thus, the bar for lethality
of oral exposure indicates adequate data (see Fig. 2.8). LOAELs and
NOAELs for target organ/systemic toxicity were available for acute,
intermediate, and chronic oral exposure of rats and mice, and data were
sufficient to derive minimal risk levels for all three durations.
Although NOAELs were not available from dog studies, LOAELs in dogs,
which were below the NOAELs in rats and mice, were used as the bases for
the intermediate and chronic minimal risk levels. Thus, the bars for
acute, intermediate, and chronic oral systemic toxicity indicate
adequate data (see Fig. 2.8). There are adequate data to indicate that
chloroform is carcinogenic in animals by the oral route but only some
data to indicate that oral exposure causes developmental effects.
Although a LOAEL and NOAEL for gonadal atrophy were available for rats,
reproductive performance was not assessed; therefore, Fig. 2.8 indicates
some data for reproductive effects. For inhalation exposure of animals,
LCsos in rats and mice were the only data on lethality; thus, the bar
indicates some data. For target organ/systemic toxicity, LOAELs, but not
NOAELs, were available for acute and intermediate inhalation exposure,
as indicated by bars for some data. There were adequate data on
developmental effects of inhalation exposure to define a LOAEL and a
NOAEL in rats and to derive a minimal risk level for acute inhalation
exposure. Some data exist for reproductive effects of inhalation
exposure, but data are lacking for carcinogenicity and systemic toxicity
of chronic inhalation exposure. The only data available for dermal
exposure of animals are that application of chloroform to the skin of
rabbits resulted in kidney effects.
2.3.2.3 Summary of relevant ongoing research
NTIS (1987) listed few studies useful for filling data gaps
identified in the preceding sections. An ongoing investigation
concerning inhalation toxicology of environmental chemicals in rats and
mice by B. Adkins of Northrop Services was identified. This study might
provide information to fill gaps in the inhalation database. A study by
W.E. Braselton of Michigan State University concerning the mechanism of
interaction between chloroform and polychlorinated biphenyls (PCBs) or
-------
Health Effects Summary 31
polybrominated blphenyls (PBBs) in causing nephrocoxicity in mice and
rats was also identified. This study appeared to be primarily concerned
with mechanisms of toxicity and metabolic pathways. NTIS (1987) also
listed a study by D.W. Deamer of the University of California-Davis
concerning anesthetic effects on membrane permeability. This in vitro
study could elucidate the mechanism of action of general anesthetics
like chloroform.
NTP (1986) plans investigations concerning the effects of time-
varying concentration profiles and patterns of repetitive inhalation
exposures to chloroform on the development of cleft palate in mice.
These studies should add to the animal inhalation developmental effects
database. NTP (1986) also plans to test chloroform in an in vivo mouse
bone marrow cytogenetic test.
The Japan Bioassay Laboratory in Kanagawa is conducting 14- and
90-day inhalation studies of chloroform in rats and mice for the
Japanese government (Davidson 1988). Chronic studies are also planned.
2.3.3 Other Information Needed for Human Health Assessment
2.3.3.1 Pharmacokinetics and mechanisms of action
The pharmacokinetics and mechanisms of chloroform toxicity in the
liver and kidneys are relatively well understood. Chloroform is
metabolized to phosgene or some other reactive metabolite, which then
causes toxic effects in liver and kidneys. The mechanism of chloroform
effects on the CNS are less well understood. The liver, kidneys, and CNS
are the primary targets of chloroform toxicity, but the relatively
limited database for inhalation and dermal exposure effects makes it
difficult to evaluate the appropriateness of route-to-route
extrapolation. The Dow Chemical Company (1988) has a joint project with
Drs. R. H. Reitz, M. E. Andersen, R. B. Connoly, and R. J. Corley to
develop a physiologically based pharmacokinetic model for chloroform,
which is anticipated to be validated and ready for use in 1988.
Additional data concerning method of administration (bolus vs continuous
dosing) and vehicle effects on pharmacokinetics, toxicity, and
carcinogenicity of chloroform are needed to evaluate the appropriateness
of available data for human health effects risk assessment.
2.3.3.2 Monitoring of human biological samples
Methods for detecting chloroform in exhaled breath, blood, urine,
and tissue are available, but monitoring data of humans for exposure to
chloroform in relation to health effects or for indicators of exposure
are inadequate due to chloroform's volatility and short half-life in
tissues. In addition, the presence of chloroform or its metabolites in
biological fluids and tissues may result from the metabolism of other
chlorinated hydrocarbons; thus, elevated tissue levels of chloroform or
its metabolites may reflect exposure to other compounds.
-------
32 Section 2
2.3.3.3 Environmental considerations
Although chloroform has been detected In some foods and beverages,
data regarding the average daily dietary intake were not located.
Limited data are available on the persistence of chloroform in the
environment, particularly in surface waters, soil, and groundwater.
Although the half-life of chloroform in surface water was estimated,
significant uncertainty exists. Due to the lack of data it was not
possible to estimate the half-life of chloroform in either soil or
groundwater.
No studies were available regarding the chemical interaction
between chloroform and other pollutants in the environment.
There are no known ongoing experimental studies intended to fill
the data gaps on the environmental fate and transport of chloroform.
-------
33
3. CHEMICAL AND PHYSICAL INFORMATION
3.1 CHEMICAL IDENTITY
Data pertaining to the chemical identity of chloroform are listed
in Table 3.1.
3.2 PHYSICAL AND CHEMICAL PROPERTIES
The physical and chemical properties of chloroform are presented in
Table 3.2.
-------
34 Section 3
Table 3.1. Chemical identity of chloroform
Value
References
Chemical name
Synonyms
Trade name
Chemical formula
Wiswesser line notation
Chemical structure
Identification numbers
CAS Registry No.
NIOSH RTECS No.
EPA Hazardous
Waste No.
OHM-TADS No.
DOT/UN/NA/IMCO
Shipping No.
STCC No.
Hazardous Substances
Data Bank No.
National Cancer
Institute No.
Trichloromethane
Methenyl chloride,
Methane trichloride,
Methyl trichloride,
Formyl trichloride
Freon 20, R 20,
R 20 (refrigerant)
CHC1)
GYGG
C1
H —C —Cl
C1
67-66-3
FS 9100000
Chemline 1987
IARC 1979
IARC 1979
NLM 1987
NLM 1987
IARC 1979
NLM, 1987
NLM 1987
U044 NLM 1987
7216639 NLM 1987
Chloroform; UN 1888 NLM 1987
49 403 10 (other than
technical grade)
49 403 11 (technical
grade)
56
C02686
NLM 1987
NLM 1987
NLM 1987
NIOSH 1987a
-------
Chemical and Physical Information 35
Table 3.2. Physical and chemical properties of chloroform
Property
Molecular weight
Color
Physical state
Odor
Odor threshold
Water
Air
Melting point. °C
Boiling point. °C
Autoignition
temperature, °C
Solubility
Water
Organic solvents
Density
Vapor density (air = 1)
Partition coefficients
Log octanol/water
Blood/air
Olive oil/air
Vapor pressure
Henry's law constant
Refractive index
Flash point
Flammability limits
Conversion factors
ppm (v/v) to mg/m1
in air (20°C)
mg/m3 to ppm (v/v)
mair(20°C)
Value
119.38
Colorless.
water-white
Liquid
Pleasant, ethereal,
nomrntating
2.4 ppm (w/v)
85 ppm (v/v)
-63.2
61 3
>1000
8.22 X I03 mg/L at 20°C
Miscible with principal
organic solvents
1 485 g/cm3 (20°C)
436
1.97
11.7
563
159mm Hg (20°C)
3.0 X 10~3 atm-mj/mol (20°C)
I.4466(20°C)
None
Unknown
ppm (v/v) - 4.96 mg/m3
mg/m3 — 0.20 ppm (v/v)
References
DeShon 1979
Hawley 1981
DeShon 1979
DeShon 1979
DeShon 1979
Amoore and Hautula
1983
Amoore and Hautula
1983
DeShon 1979
DeShon 1979
DeShon 1979
DeShon 1979
DeShon 1979
Hawley 1981
DeShon 1979
Hansch and Leo
1985
ACGIH 1986
ACGIH 1986
Boublik 1984
Nicholson et
al. 1984
DeShon 1979
DeShon 1979
-------
4. TOXICOLOGICAL DATA
4.1 OVERVIEW
Chloroform is readily absorbed through the lungs and intestinal
mucosa. Absorption from the gastrointestinal tract is rapid and complete
and follows first-order passive absorptive processes. With oral
ingestion, a dose-dependent, first-pass effect occurs with pulmonary
elimination of unchanged chloroform. Pulmonary uptake occurs by firsc-
order diffusion processes. Pulmonary uptake and elimination has three
components with rate constants corresponding to tissue loading or
desaturation of three major compartments.
Chloroform concentrations in tissues are dose related and occur in
the following order: adipose > brain > liver > kidney > blood.
The two major elimination processes for chloroform are pulmonary
elimination of unchanged chloroform and metabolism of chloroform in
liver and other tissues. Chloroform metabolism is dose dependent and
saturable. Data from in vitro and in vivo studies indicate that
chloroform is metabolized to phosgene via a cytochrome P45Q-dependent
mechanism. Phosgene or other reactive metabolites are apparently
responsible for the toxic effects of chloroform.
Because of large interspecies differences in chloroform metabolism
and marked sex differences in tissue distribution and covalent binding
to tissue macromolecules in mice, extrapolating from animal studies to
man may be difficult.
Male mice were especially sensitive to the lethal effects of
chloroform via the oral and inhalation routes. The lowest reported
effect level for acute lethality was 1025 ppm for inhalation exposures
and 118 mg/kg for oral exposures. The lowest reported lethal oral dose
for humans was 211 mg/kg.
The three principal target organs of chloroform toxicity are the
liver, the kidneys, and the CNS. Dogs appeared to be the species most
sensitive to chloroform-induced hepatotoxicity. Liver generally was the
most sensitive target organ, although the kidneys in male mice of
certain strains were the most sensitive targets. There are a large
number of oral studies but relatively few inhalation studies and almost
no dermal data for the target organ/systemic toxicity of chloroform.
Possible teratogenic effects of chloroform were observed in mice
exposed to 100 ppm in the atmosphere. Another study demonstrated
fetotoxic but no teratogenic effects in rats exposed to 30 ppm.
Developmental effects were observed in oral teratogenicity studies in
rats and rabbits. Fetotoxic effects in rabbits occurred at 20 mg/kg/day.
the lowest dose tested in any of these studies, but these effects were
not dose related.
-------
38 Section 6
There Ls little information concerning the reproductive effects of
chloroform. Experiments in animals indicate that chloroform may cause
sperm abnormalities and gonadal atrophy.
Chloroform has been tested for gene mutations in prokaryotes,
eukaryotes, cultured mammalian cells, and Drosophila, and for chromosome
aberrations in vivo and in cultured mammalian cells. No definitive
conclusions regarding the genotoxicity of chloroform can be made.
Epidemiologic studies indicate a possible relationship between
exposure to chlorinated drinking water and cancer of the bladder, large
intestine, and rectum. Chloroform is one of several volatile organic
contaminants present in chlorinated drinking water, and it has not been
identified as the sole or primary cause of excess cancer rates in this
association.
Chloroform carcinogenicity has been investigated in a number of
oral exposure animal studies. Positive results included treatment-
related increases in incidences of renal epithelial tumors in male
Osborne-Mendel rats, hepatocellular carcinomas in male and female B6C3F1
mice, kidney tumors in male ICI mice, and hepatomas in female strain A
mice. Negative results were obtained in studies with (C57 x DBA2 Fl)
mice; female Osborne-Mendel rats; female ICI mice and male mice of
strains CBA, C57BL, and CF-1; female &6C3F1 mice; male and female
Sprague-Dawley rats; and male and female beagle dogs.
4.2 TOXICOKINETICS
4.2.1 Absorption
4.2.1.1 Inhalation
Human. Respiratory absorption of chloroform ranges from 49 to 77%
(EPA 1980). According to EPA (198Sa), the amount of pulmonary chloroform
absorption is dependent on the concentration in inhaled air, the
duration of exposure, the blood/air partition coefficient, the
solubility in various tissues, and the state of physical activity, which
influences the ventilation rate and cardiac output. Pulmonary absorption
of chloroform is also influenced by total body weight and total fat
content, with uptake and storage in adipose tissue increasing with
excess body weight and obesity.
EPA (1985a) noted that in inhalation exposures, the arterial blood
concentration of chloroform is directly proportional to the
concentration in inspired air. At anesthetic concentrations (8000-10,000
ppra), Smith et al. (1973) and Morris (1951) recorded arterial blood
concentrations of 7 to 16.5 mg/dL. At lower air concentrations, blood
concentrations were proportionally lower.
Lehman and Hasegawa (1910) reported that equilibrium between blood
concentration and inhaled air concentration was reached in 80 to 100
min. Total body equilibrium with inspired chloroform concentration
requires at least 2 h in normal humans at resting ventilation and
cardiac output (Lehmann and Hasegawa 1910, Smith et al. 1973, EPA
198Sa). Using data from Smith et al. (1973) and Lehman and Hasegawa
-------
lexicological Data 39
(1910), EPA (1985a) calculated that the retention of chloroform at
equilibrium was -65%.
Animal. Although systemic toxicity observed in animals after
inhalation exposure to chloroform (see Sect. 4.3 on toxicity) indicates
that chloroform is absorbed by the lungs, data regarding the rate and
extent of absorption were not available from animal studies. Pulmonary
absorption of chloroform by animals should be similar to that of humans
4.2.1.2 Oral
Human. Chloroform absorption from the gastrointestinal tract is
-100% (Fry et al. 1972). Absorption of an oral dose of 13C-chloroform
(0.5 g in a gelatin capsule) was rapid in humans, with peak blood levels
in 1 h (Fry et al. 1972, EPA 1985a).
Animal. Experiments in mice, rats, and monkeys indicated that oraL
doses (60 mg/kg) of 14C-chloroform in olive oil were almost completely
absorbed as indicated by a 93 to 98% recovery of radioactivity in
expired air, urine, and carcass (Brown et al. 1974a, Taylor et al. 1974,
EPA 1985a).
Withey et al. (1983) compared intestinal absorption of chloroform
administered intragastrically in either water or corn oil to rats and
found that absorption was rapid with both vehicles, but that the rate
and extent of absorption varied greatly. The peak concentration of
chloroform in blood after administration in an aqueous medium was 39.3
Mg/mL when administered in water and 5.9 pg/mL when administered in corn
oil. The greater degree of absorption of chloroform in aqueous solution
compared with an oily vehicle can be explained by the faster
partitioning of a lipophilic compound such as chloroform with mucosal
lipids from an aqueous vehicle than from an oily vehicle. Peak blood
concentrations occurred slightly more rapidly with the water vehicle
(5.6 min vs 6.0 min for corn oil). The authors also noted that the
uptake from a corn oil solution was more complex (pulsed) than from
aqueous solution. They suggested that a possible explanation for this
behavior was that the chloroform in corn oil was broken up into
immiscible-globules, some of which did not contact the gastric mucosa.
Another possible explanation was that intragastric motility could have
separated the doses into aliquots that were differentially absorbed from
different sections of the gastrointestinal tract.
Absorption in mice and monkeys was rapid, with peak blood levels in
1 h following oral doses of 60 mgAg chloroform in olive oil (Brown ec
al. 1974a, Taylor et al. 1974, EPA I985a).
4.2.1.3 Dermal
Human. No experimental data concerning dermal absorption of
chloroform by humans were available. Tsurata (1975) calculated a dermal
absorption rate of 329 /imol/min/cm2 for shaved abdominal skin of mice.
equivalent to an absorption of 19.7 mg/min for human absorption if boch
hands were immersed in liquid chloroform. The calculation was based on
the assumptions that the rate of chloroform penetration is uniform for
all types of skin and that the total surface area of a pair of human
hands is 800 cm^.
-------
40 Section 4
Animal. Evidence from guinea pig dermal absorption studies with
solvents other than chloroform indicated that for solvents such as
chloroform, absorption occurs faster than metabolism or pulmonary
excretion (Jakobson et al. 1982, EPA 1985a).
Tsurata (1975, 1977) calculated a dermal absorption rate of 329
A brain > liver > kidney > blood.
Chenoweth et al. (1962) found high concentrations in fat (lOx
blood) and adrenals (4x blood) of dogs after surgical anesthesia,
whereas levels in brain, liver, and kidneys were similar to blood.
Cohen and Hood (1969) examined distribution of 14C-chloroform in
mice at 15 and 120 min after inhalation exposure and found that fat and
liver were the only organs with tissue/blood concentration ratios >2,
although kidney and liver values increased during the 2-h experiment,
indicating continuing accumulation.
Chloroform crosses the placental barrier, as indicated by
teratogenic and embryotoxic effects in mice, rats, and rabbits (Murray
et al. 1979, Dilley et al. 1977, Schwetz et al. 1974, Thompson et al.
1974) and accumulation in fetal liver (Von Oettingen 1964). Danielsson
et al. (1986) found volatile radioactivity in the placenta and fetuses
of pregnant mice a short time after inhalation of *4C-chloroform.
Chloroform metabolites accumulated with time, especially in the amniocic
fluid. In early gestation, metabolites accumulated in embryonic neural
tissues and, in late gestation, in the fetal respiratory epithelium.
Chloroform has been detected in fresh cow's milk and would be
expected to appear in human milk as well (EPA 1985a).
-------
lexicological Data -.
4.2.2.2 Oral
Human. No data concerning distribution of orally administered
chloroform in humans were found in the available literature.
Animal. Brown et al. (1974a) found high concentrations of
radioactivity in body fat and liver of rats and squirrel monkeys given
oral doses of 60 mg/kg 14C-chloroform.
Taylor et al. (1974) observed highest levels of radioactivity in
liver and kidneys of three strains of mice 3 h after oral dosing with 60
mg/kg l^C-chloroforra. Male mice had 3.5 times more activity in kidneys
than females, and this may be related to the nephrotoxicity observed IP
male mice but not females (Bennet and Whigham 1964; Culliford and Hewicr
1957; Hewitt 1956; Shubik and Ritchie 1953; Eschenbrenner and Miller
1945a,b; EPA 1985a). Sex differences in chloroform tissue distribution
were observed only in mice, not in rats or squirrel monkeys (Brown et
al. 1974a).
Cohen and Hood (1969) found higher chloroform levels in fat than IP
liver or kidneys following inhalation exposure, whereas Brown et al.
(1974a) and Taylor et al. (1974) found higher levels in liver and
kidneys than in fat following oral exposure. This suggests that
chloroform distribution in mice may depend on the route of
administration and may be related to first-pass extraction by the liver
in oral exposure, differences in observation times after exposure, and
differences in metabolism and binding of metabolites to cellular
macromolecules (EPA 1985a).
4.2.2.3 Dermal
No data concerning distribution of dermally administered chloroform
in humans or animals were found in the available literature.
4.2.3 Metabolism
A general scheme of chloroform metabolism is presented in Fig. 4 1
4.2.3.1 Inhalation
No data concerning metabolism of inhaled chloroform by humans or
animals were found in the available literature.
4.2.3.2 Oral
Human. Fry et al. (1972) found that -50% of an oral dose of 0 5 g
chloroform was metabolized to C02 by humans. The fraction metabolized
was dose dependent in that 100% of a 0.1-g dose was metabolized, but
only 35% of a 1.0-g dose was metabolized to C02•
Chiou (1975) observed that in humans as much as 38% of an oral
chloroform dose was metabolized by the liver, and up to 17% was
eliminated unchanged from the lungs before reaching the systemic
circulation, indicating a first-pass effect.
Animal. Pohl and Gillette (1984) summarized available information
concerning chloroform metabolism. Chloroform is metabolized by oxidacive
dehydrochlorination of its carbon-hydrogen bond to form phosgene. The
reaction is cytochrome P450 mediated and occurs in both liver and
-------
42
Section 6
MAJOR AEROBIC PATHWAY
P450.O2
n w^ig
ACCEPTOR
PROTEIN
1
CO
H2
NADPK
MICROSOMES
,o
//
C-CH-C-OH A
II
S NH
\ /
1 -HCI
O-CCI. H2°
PHOSGENE * 2Hd*C02
* ^^
/ CYSTE»C N.
X CONDENSATION \
^ X
QLUTATHK9NE
CONJUGATES?
H
o
2-OXOTHIAZOUDINE-
4-CARBOXYUC AGO
MINOR ANAEROBIC PATHWAY
CHCI3
ANAEROBIC
NAOPH
REDUCED
MCROSOMES
P45O-F«2<"CO
P45O-F* 2*:CCU*HCI
1
CO*2HCI
Fig. 4.1. MeUbolk pathways of chlorofonn biotraosfomutioii.
-------
Toxicological Data 43
•
kidneys. Phosgene could react with two molecules of glutathione (GSH) co
form diglutathionyl dithiocarbonate, which is further metabolized in the
kidneys, or it could react with other cellular constituents and cause
cytotoxicity. Experiments with the deuterated derivative of chloroform,
CDC13, further elucidated the mechanism of chloroform metabolism. CDC13
was 1/2 to 1/3 as cytotoxic as chloroform and was metabolized to
phosgene less rapidly than chloroform in liver and kidney tissue of rats
and mice. These findings indicated that chloroform was metabolized at
its carbon-hydrogen bond to produce toxicity and that phosgene was the
toxic agent.
Gram et al. (1986) reviewed the available information concerning
chloroform metabolism in the kidneys. In mice, chloroform is more
nephrotoxic in males than females, and the difference appears to be
mediated by testosterone. Levels of renal cytochrome P450 and associated
monooxygenases were 3 to 5 times greater in males than in females. In
addition, deuterated chloroform was less toxic than chloroform when
incubated with kidney slices from male mice. Subsequent work showed that
chloroform was metabolized to trichloromethanol, which spontaneously
dehydrochlorinates to yield phosgene. This step is rate limiting, and
since the C-D bond is stronger than the C-H bond, deuterated chloroform
is less toxic than chloroform. Phosgene may react with tissue components
or water to give C02 and HC1, or with GSH to form 2-oxothiazolidine 4-
carboxylic acid. Results of several studies (Smith and Hook 1983, 1984;
Pohl et al. 1984; Branchflower et al. 1984) indicate that the kidney
itself metabolizes chloroform to phosgene, resulting in local toxicity
The major end product of chloroform metabolism is carbon dioxide
(Brown et al. 1974a, Fry et al. 1972, Rubinstein and Kanics 1964, Van
Dyke et al. 1964, Paul and Rubinstein 1963, EPA 198Sa), most of which is
eliminated via the lungs, but some is incorporated into endogenous
metabolites and excreted as bicarbonate, urea, methionine, and other
amino acids (Brown et al. 1974a, EPA 1985a). Inorganic chloride ion is
an end product of chloroform metabolism that has been found in the urine
(Zeller 1883, Van Dyke et al. 1964, EPA 198Sa). Carbon monoxide was
found to be a minor metabolite of chloroform in in vitro (Ahmed et al
1977, Wolf et al. 1977) and in vivo studies (Anders et al. 1978, Bellar
et al. 1974, EPA 1985a).
The extent of chloroform metabolism is variable. Mice, rats, and
squirrel monkeys metabolized 85, 66, and 28% of an oral dose of 60 rag/kg
^•^C-chloroform to C02 and 2 to 8% to various urinary metabolites (Brown
et al. 1974a, Taylor et al. 1974, EPA 198Sa). In rats, 67 and 69% of an
oral dose of 12 or 36 mg/kg was metabolized to C02 (Reynolds et al.
1984a,b; EPA 1985a).
The mode of oral administration and vehicle appear to have an
effect on the metabolism of chloroform. As discussed by EPA (198Sa),
virtually all of the chloroform administered in drinking water would be
absorbed and metabolized to reactive metabolites, but a certain
percentage of a bolus gavage dose would be excreted unchanged. Using the
available pharmacokinetic data, EPA (1985a) estimated that for rats and
mice given 60 mg/kg by gavage, the amount of the dose excreted unchanged
would be 6% for mice and 20% for rats. For drinking water
administration, virtually all of the dose would be metabolized.
-------
44 Section 4
4.2.3.3 Dei
No data concerning metabolism of dermally applied chloroform in
humans or animals were found in the available literature.
4.2.4 Excretion
4.2.4.1 Inhalation
Hunan. Fry et al. (1972) observed a three-component elimination
curve for decline of blood chloroform concentration in humans. The
components were a very rapid phase with half-time of 14 min, a slower
phase with half-time of 90 min, and a very slow phase with a very long,
but undetermined, half-time.
Animal. Kinetics of pulmonary excretion of chloroform are typical
of gaseous vapor pulmonary elimination for relatively hydrophobic
volatile gaseous anesthetics and are best represented by a three-
compartment first-order model. The three compartments are vessel-rich
tissues, lean body mass, and adipose tissue (EPA 198Sa).
Because chloroform is quite soluble in body fat, it has a
relatively long half-time of elimination from this compartment, -36 h
(Stewart et al. 1965, EPA 1985a). The elimination half-time for the
vessel-rich tissues is -30 min (Lehmann and Hasegawa 1910, EPA 1985a).
4.2.4.2 Oral
Human. Fry et al. (1972) observed that 96% (radioactivity) of an
oral dose of 500 mg of 14C-labeled chloroform was exhaled as unchanged
chloroform or C02 by adult humans within 8 h, and <1% was found in the
urine. The amount of unchanged chloroform eliminated via -he lungs
within 8 h increased in proportion to dose. Lean subject eliminated a
greater percentage of the dose via the lungs than overweight subjects.
About 50% of the oral dose in the Fry et al. (1972) study was exhaled as
C02 and the rest as unchanged chloroform (Chiou 1975, EPA 1980).
Animal. Reynolds et al. (1984a,b) gave rats gavage doses of 12 or
36 mg/kg in mineral oil and found that 5 and 12%, respectively, were
excreted unchanged in exhaled air, whereas 67 and 68%, respectively,
were metabolized to C02 and exhaled. Brown et al. (1974a) and Taylor et
al. (1974) administered 60 mg/kg of 14C-chloroform orally to mice (three
strains), rats, and squirrel monkeys and found that they eliminated 6,
20, and 79%, respectively, of the administered dose via the lungs as
chloroform or metabolites other than C02- Species differences were
related to the capacity to metabolize chloroform rather than differences
in pulmonary kinetics (EPA 1985a). Humans and nonhuman primates
eliminated chloroform in the breath primarily as unchanged chloroform
(Brown et al. 1974a, Fry et al. 1972), whereas mice eliminated almost
80% of an oral chloroform dose as C02 (Taylor et al. 1974).
Mink et al. (1986) administered single oral doses of 100 mgAg
14C-chloroform to rats and mice. In rats, total recovery was 78% after
8 h, 65% of which was expired unmetabolized chloroform. Another 6.5% was
expired as C02, 2.6% was excreted in urine, and 3.6% was recovered in
organs. Chloroform was metabolized to a greater extent in mice, 26% was
-------
Toxicological Data 45
expired as unchanged chloroform and 49.6% as C02• Total recovery was
94.5%, with 13.5% in organs and 4.9% in urine.
Other routes of chloroform elimination are of minor importance
compared with pulmonary elimination. Brown et al. (1974a) found
chloroform in high concentration in bile of squirrel monkeys following
oral doses, indicating enterohepatic circulation. They found only 2, 8,
and 3% of administered radioactivity in urine and feces of monkeys.
rats, and mice, respectively, within 48 h after dosing.
4.2.4.3 Dermal
No data concerning excretion of dermally applied chloroform by
humans or animals were found in the available literature.
4.3 TOXIC ITY
4.3.1 Lethality and Decreased Longevity
4.3.1.1 Inhalation
Human. No relevant data concerning lethality of inhaled chloroform
in humans were found.
Animal. Mice exposed to 2500 ppm chloroform in the atmosphere for
2 h experienced no obvious effects, but 3100 ppm caused narcosis, and
4100 ppm was fatal. Fatal exposures were 12,300 ppm for rabbits and
16,300 for guinea pigs (Lehman and Flury 1943, Sax 1979, EPA 1985a) Von
Oettingen (1955) reported that 8000 ppm was lethal to mice within 3 h,
and 12,500 ppm was lethal within 2 h (EPA 1980). Lundberg et al. (1986)
reported a 4-h inhalation LC50 of 47.7 g/m3 (10,000 ppm) for rats, but
the cause of death was not reported.
Deringer et al. (1953) exposed C3H mice to -5 mg/L (1025 ppm)
chloroform in the air for 1 to 3 h, and some animals died within 1 day,
thus, 1025 ppm is the lowest reported lethal inhalation level. Kidney
lesions were observed in some of the mice that died.
4.3.1.2 Oral
Human. Based on case reports, the mean lethal oral dose for humans
was estimated at -44 g (Gosselin et al. 1976). A fatal dose may be as
little as 10 mL (14.8 g or 211 mg/kg for a 70-kg human), with death due
to respiratory or cardiac arrest (Schroeder 1965, EPA 1985a).
Animal. Reported oral U>50 values are 444 to 2000 mg/kg for rats
and 118 to 1400 mg/kg for mice. Kimura et al. (1971) reported the
following oral LD50 values for rats: 0.3 mLAg (444 mg/kg) for 14-day -
old rats, 0.9 mLAg (1332 mgAg) for young adults, and 0.8 mLAg (1184
mg/kg) for older adults. Smyth et al. (1962) reported an oral LD50 of
2.18 mLAg (3226 mgAg) for rats. Chu et al. (1980) reported oral LD50
values of 908 and 1117 mgAg in male and female rats, respectively.
Torkelson et al. (1976) reported an oral LD50 of 2.0 g/kg for male rats
Liver and kidney lesions were observed. Bowman et al. (1978) reported
oral LDsos of 1120 mgAg and 1400 mgAg in male and female mice,
respectively. Mice that died had fatty infiltration of the liver and
signs of hemorrhage in the kidney, adrenal, lungs, and brain.
-------
46 Section 4
Hill (1978) administered single oral doses of chloroform to male
mice of three different strains. LDso values were 0.08, 0.20, and 0 33
raL/kg (118, 296, and 488 mg/kg) for DBA/2J, (B6D2F1)/J, and C57BL/6J
mice, respectively. Thus, 118 mg/kg is the lowest reported acute oral
LD50- Mice that died had liver and kidney lesions.
Hjelle et al. (1982) reported that a single oral dose of 60 mg/kg
had no toxic effects on mice.
Two oral studies of intermediate duration reported increased
mortality. Roe et al. (1979) administered gavage doses of 0, 60, 150, or
425 rag/kg/day chloroform in toothpaste to groups of 10 mice/sex, 6
days/week for 6 weeks. All high-dose mice died, and 8 of 10 males at 150
mg/kg died. The cause of death was not reported. Weight gain of females
was markedly reduced at 150 mg/kg, and weight gain of both sexes was
moderately retarded at 60 mg/kg. No other effects were reported. Chu et
al. (1982b) exposed groups of 20 rats/sex to 0, 5, 50, 500, or 2500 ppm
chloroform in drinking water for 90 days. High-dose rats experienced
increased mortality. The rats were emaciated prior to death.
Histological examination revealed atrophy of the liver and squamous
debris in the esophagus and gastric cardia, indicating that starvation
was the cause of death.
In a chronic study, NCI (1976) found a dose-related decrease in
survival at gavage doses of 90 or 180 mg/kg/day (males) or 100 or 200
nig/kg/day (females) administered in corn oil, 5 days/week, to OM rats
(50/sex/group) for 78 weeks followed by a 33-week observation period.
NCI (1976) also administered gavage doses of 0, 138, or 227
mg/kg/day (males) or 0, 238, or 448 mg/kg/day (females) to B6C3F1 mice
(50/sex/group) 5 days/week for 78 weeks followed by 14 to 15 weeks of
observation. Survival of high-dose females was decreased relative to
controls, but the cause of death was not reported.
Heywood et al. (1979) administered 0, 15, or 30 mg/kg/day
chloroform in toothpaste in gelatin capsules to beagle dogs
(8/sex/group) 6 days/week for 7.5 years followed by 20 to 24 weeks of
observation. No effects on survival, growth, organ weights,
hematological parameters, or urine chemistry were observed.
4.3.1.3 Dermal
Pertinent data regarding lethality following dermal exposure of
humans or animals to chloroform were not located in the available
literature.
4.3.2 Systemic/Target Organ Toxlcity
4.3.2.1 Liver effects
Overview. Liver effects were observed in humans occupationally
exposed to chloroform. Effects on the liver have been observed in rats.
mice, and dogs in inhalation and oral studies regardless of the duration
of exposure. Dogs appeared to be more sensitive to the hepatotoxic
effects of chloroform in intermediate and long-term oral exposures than
rats or mice. The hepatotoxic effects of chloroform appear to be
mediated by a reactive metabolite which can bind to microsomal protein.
-------
Toxicologies! Data 67
Inhalation, human. Challen et al. (1958) reported that 9 of 10
workers occupationally exposed to chloroform vapors at breathing zone
concentrations of 77 to 237 ppm experienced various symptoms including
thirst, irritability, lassitude, and frequent and burning urination.
Eight of ten employees exposed to 22 to 71 ppm complained of less severe
symptoms. No evidence of liver damage was found in any of these
employees who submitted to liver function tests and medical
examinations.
Bomski et al. (1967) examined liver size and function in 294
workers in a division of a pharmaceutical plant where chloroform was
used as a main solvent. The concentration of chloroform in the air of
the production area was 0.01 to 1.0 mg/L (2 to 205 ppm), whereas other
solvents were present in only trace amounts. The groups consisted of 68
people who had worked at the plant for 1 to 4 years and still had
contact with chloroform; 39 people with previous exposure to chloroform;
23 people with a history of viral hepatitis and jaundice, but with no
exposure; and 165 people with no viral hepatitis and no chloroform
exposure. Higher incidences of enlarged livers were found in the
chloroform-exposed groups (25% in those still exposed, 12.8% in those
with previous exposure) than in the controls (8.7% in the posthepatitis
nonexposed group; 3.7% in the nonhepatitis, nonexposed group). Of those
still exposed who had enlarged livers, 5.9% had toxic hepatitis
(diagnosed on the basis of increased SCOT, SGPT, and serum gamma
globulin), and 82% had fatty livers. The cases of liver disease in the
group still exposed improved when hygienic conditions improved. Phoon
et al. (1983) reported cases of toxic jaundice among factory workers
occupationally exposed to chloroform. Air analysis revealed that
concentrations were 14.4 to >400 ppm. The lack of suitable controls and
the presence of complicating factors make these studies unsuitable for
risk assessment (EPA 1985a).
Inhalation, animal. Liver effects were also observed in animals
exposed acutely by inhalation. Fatty infiltration of the liver was found
in mice exposed to 100 ppm for 4 h, and dose-related necrosis occurred
at 200, 400, and 800 ppm (Kyiin et al. 1963). Lundberg et al. (1986)
reported a 4-h TC50 of 585 mg/m3 (120 ppm) for liver damage in rats. In
a study of intermediate duration, Torkelson et al. (1976) exposed
several species to 25, 50, or 85 ppm chloroform in the atmosphere
7 h/day, 5 days/week for 6 months, or to 25 ppm 1 to 4 h/day for 6
months. Exposure to 25 ppm chloroform for up to 4 h/day for 6 months had
no adverse effects on male rats, as indicated by organ and body weights
and gross and histological examination of liver and kidneys. Exposure co
the same concentration for 7 h/day, however, resulted in
histopathological changes in the liver (lobular granular degeneration
and focal necrosis) of rats, guinea pigs, and rabbits. More severe
changes were seen in rats exposed to 50 or 85 ppm for 7 h/day, but not
in guinea pigs and rabbits.
Oral, human. Wallace (1950) described a patient who ingested 1.6
to 2.6 g chloroform/day in cough medicine for -10 years. Clinical tests
indicated that he suffered from hepatitis and nephrosis. Although
Wallace (1950) attributed these conditions to the patient's ingestion of
chloroform, there were other materials present in the cough medicine,
-------
48 Section 4
and the patient had chronically ingested moderate amounts of alcohol
daily until about 1 year prior to the examination.
Chloroform is acutely toxic to the liver, although the damage may
not be fully apparent until 24 to 48 h after exposure. Effects include
centrilobular necrosis and reduced prothrombin formation (Goodman and
Oilman 1980, Wood-Smith and Stewart 1964, EPA 1985a).
DeSalva et al. (1975) investigated the safety of a dentifrice
containing 3.4% chloroform and a mouthwash containing 0.43% chloroform
in two studies lasting 1 and 5 years and involving 229 subjects. Persons
using the dentifrice were exposed to -68 mg/day, while those using the
mouthwash and toothpaste were exposed to -197 mg/day. There were no
differences in liver function tests (SGPT, SCOT, and SAP) between
experimental and control subjects. The authors assumed that the average
ingestion of the dentifrice and mouthwash was 25%, resulting in
ingestion of 0.34 mg/kg in the first study and 0.96 rag/kg in the second
study, using an assumed adult weight of 50 kg.
Oral, animal. Three short-term oral studies in mice found liver
effects. Dose-related increased incidence and severity of centrilobular
cytoplasmic pallor, mitotic figures, and focal inflammation in the liver
were observed in mice administered single gavage doses of 37, 74, or 148
mgAg chloroform in corn oil compared with controls (Condie et al.
1983). SGPT levels were elevated at the highest dose. Jones et al.
(1958) observed fatty infiltration in the liver of mice after a single
dose of 30 mg/kg. and doses of 133 to 355 mgAg caused hepatic damage.
including centrilobular necrosis. These doses represent an acute oral
LOAEL and PEL, respectively, for hepatic effects in mice. Moore et al.
(1982) found increased SCOT and uptake of thymidine in the livers of
mice after single oral doses of chloroform in corn oil at 273 mg/kg or
in toothpaste at 199 mgAg. Doses of -18 and -60 mg/kg in either vehicle
did not have any toxic effects on the liver. Thus, 18 mgAg appears to
be a NOAEL for short-term oral exposure in mice.
Munson et al. (1982) administered aqueous gavage doses of 0, 50,
125, and 250 mgAg/day to groups of 7 to 12 male and 8 to 12 female CD-I
mice for 14 consecutive days. Liver weights were increased at all doses
in females and at 125 and 250 mgAg/day in males. SCOT and SGPT were
increased in high-dose females, and serum glucose was decreased at the
intermediate and high doses. No effects were reported at 50 mgAg/day.
For short-term exposure of rats, a teratology study (Sect. 4.3.3.2)
provided dose-response data for liver effects. Thompson et al. (1974)
observed fatty infiltration at 50 mgAg/day (LOAEL) and hepatitis at 316
mgAg/day (FEL), but no effects at 20 mgAg/day (NOAEL) were observed in
pregnant rats treated by corn oil gavage for 10 days during gestation.
Torkelson et al. (1976) observed gross pathologic changes in the liver
of rats at acute oral doses of 2250 mgAg- Tumasonis et al. (1985),
however, reported that exposure to chloroform in drinking water at 2900
mg/L (406 mgAg/day) for 3 weeks did not cause histopathological lesions
in the liver of rats. The difference in response in these studies is
probably due to the different mode of administration (i.e., gavage vs
drinking water).
-------
ToxicoLogical Data i9
There are several subchronic oral studies chat provide dose-
response data for liver effects in rats, mice, and dogs.
Palmer et al. (1979) administered gavage doses of 0, 15. 30. 150,
or 410 mgAg/day in toothpaste, 6 days/week, to SD rats (10/sex/group)
for 13 weeks. High-dose rats experienced increased liver weight with
fatty change and necrosis. Relative liver weight was affected at 150
nag/kg, but no effects occurred at lower doses. For intermediate duration
in rats. 410 mg/kg/day is a PEL. 150 rag/kg/day is a LOAEL, and 30
mg/kg/day is a NOAEL for liver effects. Subchronic drinking water
studies report milder liver effects at levels equivalent to higher
doses. In the 90-day drinking water study by Chu et al. (1982b) (see
Sect. 4.3.1.2), rats exposed to 2500 ppm (350 mg/kg/day). but not <500
ppm (70 mg/kg/day), had increased frequency of mild to moderate liver
lesions compared with controls. The severity of these lesions, however,
was similar to those seen in controls. Rats treated with drinking water
containing chloroform at 200 to 1800 ppm (28-252 mg/kg/day) for 90 days
had increased incidences of "hepatosis" compared with controls, but the
incidence was not dose related (Jorgensen and Rushbrook 1980). The
equivalent doses were calculated assuming that a 0.35-kg rat consumes
0.049 L water per day (EPA 1985b).
Bull et al. (1986) administered gavage doses of 0, 60, 130, or 270
mg/kg/day chloroform in corn oil (or 2% emulphor suspension) to groups
of 10 B6C3F1 mice/sex for 91 to 94 days. Chloroform administered in corn
oil had greater hepatotoxic effects than chloroform administered in the
aqueous suspension. With corn oil as the vehicle, there were increased
liver lipid levels and extensive vacuolation of hepatocytes at 60
nig/kg/day. There was also extensive disruption of the normal hepatic
architecture accompanied by infiltration of inflammatory and spindle
cells at 270 mg/kg/day. Pathology in mice treated with chloroform in 2%
emulphor consisted of minimal to mild focal necrosis at 130 and 270
nig/kg/day.
Jorgenson and Rushbrook (1980) exposed groups of 30 female B6C3F1
mice to 0, 200, 400, 600, 900, 1800, or 2700 ppm (0, 40, 80, 120, 180,
360, or 540 mgAg/day) chloroform in drinking water for up to 90 days.
There was mild hepatic centrilobular fatty change in all treated groups
except the 200-ppm (40 mgAg/day) group at 30 days, but only in the
1800- and 2700-ppm dose groups at 60 and 90 days.
Munson et al. (1982) administered chloroform as an aqueous solution
in emulphor at doses of 0, 50, 125, or 250 mgAg/day to groups of 7 to
12 male and 7 to 12 female CD-I mice for 90 days. Liver weights and
serum glucose levels were increased in high-dose males. Liver weights
were increased in all treated females. Treated mice also exhibited
slight histopathologic changes in liver and kidneys that were not seen
in controls. Kidneys of treated mice had small intertubular collections
of chronic inflammatory cells, mostly lymphocytes. In the liver there
was generalized hydropic degeneration of hepatocytes and small focal
collections of lymphocytes. The lowest dose used in this study, 50
mgAg/day, represents a LOAEL for liver effects in mice.
Heywood et al. (1979) administered 0, 30, 45, 60, 90, or 120
mgAg/day chloroform in gelatin capsules, 7 days/week, for up to 18
weeks to groups of one to two beagle dogs/sex/dose. Frequently, elevated
-------
SO Section 4
SGPT, SAP, and SCOT and hepatocyte enlargement with vacuolization and
fat deposition were observed at >60 mg/kg/day. Increased relative liver
weight, discoloration, and variation in hepatocyte size (with slight fat
deposition) occurred at 45 mg/kg. SGPT was occasionally elevated at 45
and 30 mg/kg/day. No histopathological effects were observed at 30
mg/kg. The lowest dose used in this study, 30 mg/kg/day, is considered
an intermediate-duration LOAEL for liver effects in dogs because of
occasionally elevated SGPT.
Several chronic studies also found effects on the livers of rats,
mice, and dogs.
NCI (1976) administered gavage doses of 0, 90. or 180 mg/kg/day
(males) or 0, 100, or 200 mg/kg/day (females) in corn oil, 5 days/week,
to OM rats (50/sex/group) for 78 weeks followed by a 33-week observation
period. Increased incidences of necrosis of hepatic parenchyma in males
and females were dose related. Palmer et al. (1979) conducted a study
with groups of 50 rats/sex receiving gavage doses of 0 or 60 mg/kg/day
in toothpaste, 6 days/week, for 80 weeks followed by 15 weeks of
observation. There were minor histological changes in livers and
decreased relative liver weights in female rats, but there was no
evidence of chloroform-induced hepatotoxicity. Thus, 60 mg/kg/day
appears to be a chronic oral NOAEL, and 90 mg/kg/day a LOAEL for liver
effects in rats.
NCI (1976) administered gavage doses of 0, 138, or 227 mgAg/day
(males) or 0, 238, or 448 mgAg/day (females) chloroform in corn oil to
groups of 50 B6C3F1 mice/sex, 5 days/week for 78 weeks followed by 14 to
15 weeks of observation. Nonneoplastic proliferative changes and
necrosis occurred in livers of both sexes at both doses.
Heywood et al. (1979) administered 0, 15, or 30 mg/kg/day
chloroform in toothpaste in gelatin capsules to groups of beagle dogs
(8/sex/group), 6 days/week for 7.5 years followed by 20 to 24 weeks of
observation. No effects on survival, growth, organ weights,
hematological parameters, or urine chemistry were observed. There was a
moderate dose-related increase in SGPT and other serum enzymes
indicative of liver damage, which reached a peak in the sixth year of
the study but reverted to normal levels after treatment ended.
Aggregation of vacuolated histiocytes ("fatty cysts") occurred in all
groups; however, they were larger and more numerous in treated dogs.
Therefore, an oral dosage of 15 mgAg/day may be considered a LOAEL for
liver effects in dogs.
Dermal. No dermal toxicity studies showing effects of chloroform
on the liver of humans or animals were found in the available
literature.
General discussion. Dogs appear to be the animal species most
sensitive to the hepatotoxic effects of chloroform. Reported adverse
effect levels for liver effects in dogs were below NCAELs for rats and
mice in chronic studies.
In vitro studies indicated that phosgene and other reactive
chloroform metabolites bind preferentially to lipids and proteins of the
endoplasmic reticulum proximate to the P450 metabolic system. Covalent
binding also occurs in other cell fractions of the liver and kidneys,
-------
Toxicological Data 51
especially to mitochondria (Uehleke and Werner 1975, Hill et al. 1975
EPA 1985a).
Covalent binding of chloroform metabolites to microsomal protein in
vitro was enhanced by microsomal enzyme inducers and prevented by
glutathione. Brown et al. (1974b) proposed a mechanism of toxicity
involving formation of a free radical metabolite that can react with
glutathione, and, after glutathione is depleted, with microsomal
protein, causing necrosis. Docks and Krishna (1976) found that only
those doses of chloroform causing liver glutathione depletion caused
liver necrosis. Ekstrom et al. (1982) reported that chloroform or a
reactive metabolite inhibited glutathione synthesis. Chlorofsrm
hepatotoxicity apparently depends on (1) the rate of its transformation
to active metabolites and (2) the amount of glutathione available to
conjugate and inactivate metabolites (EPA 1985a). Chloroform is more
hepatotoxic in fasted than in fed animals, possibly due to decreased
hepatic glutathione content in fasted animals (Docks and Krishna 1976,
Brown et al. 1974b, EPA 1985a).
4.3.2.2 Kidney effects
Overview. Male mice of certain strains are especially sensitive to
toxic effects of chloroform on kidneys. Kidney lesions have been
observed in mice after short-term inhalation exposures and after short-
term, intermediate, and long-term oral exposures. Strain differences in
susceptibility to kidney effects from chloroform appear to be related to
the ability of the kidney to metabolize chloroform to phosgene. Sex
differences may be related to testosterone levels. Subchronic inhalation
exposures and short-term and subchronic oral exposures caused adverse
effects in kidneys of rats. Kidney lesions were also observed in rabbits
after short-term dermal exposure. Evidence that chloroform has effects
on the kidneys of humans is sparse.
Inhalation, human. No inhalation data concerning chloroform
effects on human kidneys were found in the available literature.
Inhalation, animal. Deringer et al. (1953) exposed C3H mice to
-5 mg/L (1025 ppm) chloroform in the air for 1 to 3 h and observed
kidney lesions in all of the males but none of the females. Some animals
died within 1 day and exhibited necrosis of parts of the proximal and
distal tubules.
In a study of intermediate duration, Torkelson et al. (1976)
exposed several species to 25, 50, or 85 ppm chloroform in the
atmosphere, 4 or 7 h/day, 5 days/week for 6 months. Exposure to 25 ppm
chloroform for up to 4 h/day for 6 months had no adverse effects on male
rats, as indicated by organ and body weights and gross and histological
examination of liver and kidneys. Exposure to the same concentration for
7 h/day, however, resulted in cloudy swelling of the kidneys of rats,
guinea pigs, and rabbits. More severe changes were seen in rats (but noc
in guinea pigs and rabbits) exposed to 50 or 85 ppm for 7 h/day.
Oral, human. Wallace (1950) described a patient who ingested 1.6
to 2.6 g chloroform per day in cough medicine for -10 years. Clinical
tests indicated that he suffered from hepatitis and nephrosis. Although
Wallace (1950) attributed these conditions to the patient's ingestion of
-------
52 Section 4
chloroform, there were other materials present in the cough medicine,
and the patient had chronically ingested moderate amounts of alcohol
daily until about 1 year prior to the examination.
Oral, animal. Kidney lesions have been observed in mice after
single oral doses. Moore et al. (1982) administered single oral doses of
0, 17.3, 65.6, or 273 mg/kg chloroform in corn oil or 0, 18.2, 59.2. or
199 mgAg in toothpaste to groups of 3 to 5 male CFLP outbred Swiss
albino mice. The low dose of -18 mg/kg chloroform in either vehicle did
not have any toxic effects on kidneys of mice or stimulate any
regenerative activity. Doses of -60 mg/kg caused an increase in kidney
weight, tubular necrosis, and areas of basophilia, indicating tubular
regeneration. The high doses caused kidney necrosis, increased thymidine
uptake, and elevated plasma urea in all animals. Thus, an acute oral
dose of 18 mgAg is a NOAEL for kidney effects in mice. The severity of
toxic effects was greater when chloroform was administered in corn oil
rather than toothpaste.
These dose -response data are supported by other short-term oral
studies in mice. Reitz et al. (1980) observed severe diffuse renal
necrosis in male mice after a single oral dose of 240 mg/kg and focal
tubular regeneration after 60 or 240 mg/kg, but no effects at 15 mg/kg.
Eschenbrenner and Miller (1945b) also observed renal necrosis in male
(but not female) strain A mice given single intragastric doses of
>148 mgAg chloroform. Condie et al. (1983) administered gavage doses of
0, 37, 74, or 148 mg/kg chloroform in corn oil to male CD-I mice for 14
consecutive days. Renal cortical slice uptake of PAH was decreased at
the two highest doses, and blood, urea, and nitrogen levels were
elevated at the highest dose. Dose-related increased incidence and
severity of renal intratubular mineralization, epithelial hyperplasia,
and cytomegaly were observed in kidneys of treated mice at >37
mg/kg/day. Thus, 37 mg/kg/day is the LOAEL for kidney effects for
short-term oral exposure in mice.
In rats, Chu et al. (1982a) observed increased kidney weights and
hematological and biochemical changes after a single oral dose of 1071
ngAg, but not 756 or 546 mgAg. Torkelson et al. (1976), however,
observed gross pathologic changes in kidneys of rats after single oral
doses S250 mgAg. Tumasonis et al. (1985) reported that exposure to 2900
mg/L (406 mgAg/day) chloroform in drinking water for 3 weeks did not
cause histopathologic lesions in kidneys of rats. Again, the difference
in response is probably due to the different mode of administration
(i.e., gavage vs drinking water). Rats exposed to 5, 50, or 500 ppm
(0.7, 7, or 70 mgAg/day) chloroform in the drinking water for 28 days
did not have kidney lesions (Chu et al. 1982a) . Palmer et al. (1979)
reported increased relative kidney weights in rats treated by gavage
with chloroform in toothpaste, 6 days/week for 13 weeks at 150
, but not at 230 mgAg/day.
A chronic oral study also reported effects on the kidneys of mice.
Roe et al. (1979) administered gavage doses of 0, 17, or 60 mg/kg/day
chloroform in toothpaste or arachis oil to different strains of mice, 6
days/week for 80 weeks followed by 16 to 24 weeks of observation. There
were no treatment- related effects on hematological parameters or on any
tissue except kidneys. There was an increased incidence of moderate to
-------
ToxicologicaL Data 52
severe renal changes, kidney disease, and benign and malignant tumors IT.
groups treated with 60 mg/kg/day. The tumor effect was more pronounced
when chloroform was administered in arachis oil. No adverse effects
occurred in the 17 mg/kg/day group. Thus, 60 mg/kg/day is a chronic oral
FEL for kidney effects in mice, and 17 mg/kg/day is a NOAEL.
Dermal, human. No data concerning the toxic effects of dermal
exposure to chloroform on human kidneys were found in the available
literature.
Dermal, animal. Dermal applications as low as 1000 mg/kg for 24 h
caused degenerative changes in kidney tubules of rabbits. Dermal effects
consisted of slight hyperemia with moderate necrosis (Torkelson et al
1976).
General discussion. Hook and Smith (1985) reviewed the available
information concerning the mechanism of chloroform-induced renal
toxicity. They concluded that chloroform was bioactivated in the kidney
by cytochrome P450 to a metabolite that bound irreversibly to protein
This binding could be diminished by GSH or enhanced by phenobarbital.
The identification of a phosgene-cysteine conjugate, 2-oxothizaolidine-
4-carboxylic acid, indicated that phosgene was the reactive metabolite
produced in the kidney. Bailie et al. (1984) performed several in vitro
experiments with rabbit kidneys that indicated that chloroform was
metabolized to phosgene.
Male mice of certain strains appear to be more sensitive to
chloroform-induced renal toxicity than other experimental animals. Pohl
et al. (1984) found that sensitivity to chloroform-induced renal
toxicity in mice was closely related to the capacity of the kidney to
metabolize chloroform. Kidney homogenates from male mice metabolized
chloroform almost an order of magnitude more rapidly than females.
Kidney homogenates from males of a sensitive strain metabolized
chloroform more rapidly than those from a less sensitive strain.
Renal toxicity of chloroform is probably due to its metabolism in
the kidney itself rather than the transport of toxic metabolites from
the liver (EPA 1985a). McMartin et al. (1981) reported data indicating
that the chloroform metabolite(s) responsible for hepatic and renal
toxicity are produced in those organs themselves rather than produced
elsewhere and transported to liver and kidneys.
Eschenbrenner and Miller (1945b) observed extensive renal tubular
necrosis in normal male mice and testosterone-treated castrated male
mice after oral administration of chloroform; however, necrosis was noc
observed in normal female mice or nontreated castrated male mice.
Culliford and Hewitt (1957) also investigated sex and strain differences
in chloroform renal toxicity. Adult male CBA and UH mice experienced
extensive tubular necrosis after inhalation exposure to high
concentrations for 2 h; females did not experience necrosis. Treatment
with androgens rendered females susceptible to renal necrosis, and
estrogen treatment reduced susceptibility of males. Castration
eliminated the susceptibility of one strain but not the other, however,
susceptibility of the second strain was eliminated when castration was
followed by adrenalectomy. In contrast, liver damage also occurred in
almost all exposed mice but was not correlated with sexual hormone
-------
54 Section 4
levels. Hill (1978) found that most sensitive strains of mice also
accumulated the most radioactivity in kidneys when radiolabeled
chloroform was administered. Strain and sex differences were related to
testosterone in that females and testosterone-deficient strains were
less sensitive to renal toxicity. Testosterone may sensitize renal
tubules to chloroform toxicity via a testosterone receptor mechanism
(Hill 1978) or by inducing changes in kidney morphology and physiology
(Eschenbrenner and Miller 1945b).
Smith et al. (1984) also investigated hormonal effects on
chloroform-induced nephrotoxicity in mice. They found that differences
in sensitivity were apparently related to renal MFO activity. Castration
of males reduced sensitivity and renal cytochrome P450 to levels similar
to those in females. Testosterone treatment increased chloroform
nephrotoxicity and renal cytochrome P45Q levels in both sexes.
4.3.2.3 CNS effects
Overview. The effects of chloroform on the CNS have been well
documented as a result of the use of chloroform as a surgical
anesthetic. Concentrations required for the induction of anesthesia are
20,000 to 40,000 ppm. Effects at lower levels include dizziness,
vertigo, and headaches at concentrations of -1000 to 1500 ppm.
Inhalation, human. Chloroform inhalation has a depressant effect
on the CNS, with concentrations of 20,000 to 40,000 ppm used to induce
anesthesia and lower concentrations to maintain it (NIOSH 1974, Adriani
1970, EPA 1985a). According to Goodman and Oilman (1980), a
concentration of <1500 ppm is insufficient to produce anesthesia; 1500
to 2000 ppm results in light anesthesia after induction; and 2000 to
15,000 ppm results in deep anesthesia. Chloroform anesthesia also
sensitizes the heart to epinephrine, causing arrhythmias (Kurtz et al
1936, Orth et al. 1951, EPA 1985a). Delayed toxic effects of chloroform
after use as an anesthetic included drowsiness, restlessness, jaundice,
vomiting, fever, elevated pulse rate, liver enlargement, abdominal
tenderness, delirium, coma, and abnormal liver and kidney function
(EPA 1980).
Lehman and Hasegawa (1910) and Lehman and Schmidt-Kehl (1936)
conducted inhalation experiments with humans exposed to various
chloroform concentrations for up to 30 min. They reported dizziness and
vertigo after exposure to 920 ppm for 3 min, and headache and slight
intoxication at higher concentrations. No symptoms were reported by
subjects exposed to 390 ppm for 30 min (EPA 1985a). Exposures of <30
minutes are insufficient to achieve pulmonary steady state, but longer
exposures to these concentrations might result in more severe effects
(EPA 1985a).
Challen et al. (1958) investigated health effects of occupational
exposure to chloroform among workers in a factory making medicinal
lozenges containing chloroform. Nine of ten workers exposed to -77 to
237 ppm complained of tiredness, depression, irritability, thirst,
gastrointestinal distress, and frequent and burning urination. Eight of
ten workers exposed to 22 to 71 ppm experienced less severe symptoms.
Both groups had been exposed to occasional peak concentrations of -1163
ppm for 1 to 2 min. None of five controls experienced these symptoms.
-------
lexicological Data 55
NIOSH (1974) reported that a 33-year-old male who habitually
Inhaled chloroform for 12 years experienced psychiatric and neurologic
signs of depression, loss of appetite, hallucination, ataxia, and
dysarthria.
Inhalation, animal. EPA (1985a) summarized data concerning CNS
effects of chloroform inhalation in animals. In mice, 2500 ppm for 2 h
caused no obvious effects; 3100 ppm for 1 h caused slight narcosis; and
4000 ppm caused deep narcosis in 30 min. Cats exposed to 7200 ppm
experienced disturbed equilibrium after 5 min and narcosis as exposure
duration increased.
Oral, human. No data concerning human CNS effects of oral exposure
to chloroform were located in the available literature.
Oral, animal. Jones et al. (1958) reported that a single oral dose
of 350 mg/kg was the minimum narcotic dose for rats.
Jorgenson and Rushbrook (1980) reported dose-related signs of
depression only during the first week of exposure in rats receiving oral
chloroform doses of 20 to 290 mg/kg/day for 90 days.
Balster and Borzelleca (1982) found that gavage doses of 3.1 or
31.1 /ig/kg/day chloroform had no effects on behavior of mice exposed for
14 or 90 days. Administration of 100 or 400 mg/kg/day for 60 days
affected operant behavior, but 100 mg/kg/day for 30 days did not affect
learning.
Dermal, human. Dermal exposure may contribute to the CNS
depression observed in humans occupationally exposed to chloroform.
Dermal, animal. No animal data concerning CNS effects of dermally
applied chloroform were located in the available literature.
General discussion. CNS effects of chloroform are well documented
Harris and Groh (1985) found that the disordering effect of chloroform
and other anesthetics on membrane lipids was enhanced by gangliosides.
They suggested that this might explain why the outer leaflet of the
lipid bilayer of neuronal membranes, which has a large ganglioside
content, is unusually sensitive to anesthetic agents. A study by Veiro
and Hunt (1985) indicated that anesthetics inhibited membrane
permeability independently of the channel system or type of lipid used,
suggesting that hydrogen-bonded water structure and/or hydrogen-bonding
centers at dipolar lipid-polypeptide interfaces could be sites of action
of general anesthetics. Caldwell and Harris (1985) suggested that
anesthetics affect calcium-dependent potassium conductance in the CNS.
Some of these mechanisms may contribute to effects of chloroform on the
CNS.
4.3.3 Developmental Toxicity
4.3.3.1 Inhalation
Human. No data concerning human developmental effects of inhaled
chloroform were found in the available literature.
Animal. Schwetz et al. (1974) found that chloroform was a
developmental toxicant in rats. They exposed pregnant SD rats to 0, 30,
-------
56 Section 4
100, or 300 ppm chloroform. 7 h/day, on gestation days 6 to 15.
Increased incidences of missing ribs, short or missing tail, imperforate
anus, subcutaneous edema, and delayed ossification of sternebrae
occurred at 100 ppm. Subcutaneous edema and abnormalities of the skull
and sternum occurred at 300 ppm, but their incidence was not
statistically significant, possibly because of the small number of
surviving fetuses (average of 4 per litter vs 10 per litter for
controls). Exposure to 300 ppm also caused a decrease in pregnancy rate,
number of live fetuses per litter, and an increased percentage of
litters with absorptions. Maternal weight gain was depressed at all
doses. Thus, 100 and 300 ppm are inhalation FELs, and 30 ppm is a NOAEL
for developmental toxicity in rats. Dilley et al. (1977), however,
reported in abstract form that exposure to 20 g/m^ (-4000 ppm) on days 7
to 14 of gestation caused increased fetal mortality and decreased fetal
weight gain, but no malformations in rats.
Murray et al. (1979) exposed groups of 35 to 40 pregnant CF-1 mice
to 100 ppm chloroform for 7 h/day on days 1 to 7, 6 to 15 or 8 to 15 of
gestation. There was a significant increase in the number of resorptions
per litter in animals exposed on days 1 to 7. Mean fetal body weight and
crown to rump length were significantly decreased in mice exposed on
days 1 to 7 and 8 to 15. Maternal weight gain and food and water
consumption were depressed in all groups. Incidence of cleft palates was
increased only in mice exposed on days 8 to 15.
4.3.3.2 Oral
Human. No data concerning human developmental effects of oral
exposure to chloroform were found in the available literature.
Animal. Thompson et al. (1974) administered oral doses of
chloroform to pregnant SD rats on days 6 to 15 of gestation and to
pregnant Dutch-Belted rabbits on days 6 to 18 of gestation. In the
initial range - finding experiment with rats, doses of 0, 79, 126, 300,
316, or 516 mg/kg/day chloroform in corn oil were used. Food consumption
and body weight gains of dams were significantly decreased at >126
mgAg/day. One of six rats at 316 mg/kg/day and four of six rats at 516
n»g/kg/day died and had pathologic findings of acute toxic nephrosis and
hepatitis. Fetal development was affected only at 316 and 501 mg/kg/day
where there was a significant increase in resorption rate and a decrease
in viable litter size and fetal body weights. No external abnormalities
were observed in surviving fetuses. In the main study, doses of 0, 20,
50, or 126 mg/kg/day were administered to groups of 25 pregnant rats. No
adverse effects occurred at 20 mg/kg/day, but maternal toxicity
(decreased body weight gains, mild fatty change in the liver) was
evident at i50 mg/kg/day, and fetal body weights were significantly
decreased in the 126-mg/kg/day group. No treatment-related malformations
were seen at any dose. The significance of the decreased fetal body
weights at 126 mg/kg/day is debatable because the decreases occurred at
a maternally toxic dose. In discussing the significance of such effects,
EPA (1986b) stated that "a change in offspring body weight is a
sensitive indicator of developmental toxicity," and that "current
information is inadequate to assume that developmental effects at
maternally toxic doses result only from the maternal toxicity." For
-------
lexicological Data 57
these reasons, it is not possible in this study to define a NOAEL and
LOAEL for developmental toxicity in rats.
In the range-finding study with rabbits, doses of 0, 25, 63, 100,
159, 251, or 398 mg/kg/day were given to groups of five pregnant
rabbits. Maternal toxicity occurred at >63 mg/kg/day. In the main study,
groups of 15 pregnant rabbits were given 0, 20, 35, or 50-mg/kg/day.
Maternal weight gain was depressed at 50 mg/kg/day, and mean fetal body
weight was depressed at 20 and 50 but not 35 mg/kg/day. No teratogenic
effects were observed. Incidence of incompletely ossified skull bones
was significantly increased among fetuses, but not litters, at 20 and 35
mg/kg/day but not at 50 mg/kg/day. Because the incidences of incomplete
ossification were not dose related, and because the incidences using the
litter rather than the fetus as the unit of comparison were not
statistically significant, the 50-mg/kg/day dose can be considered a
NOAEL for developmental toxicity in rabbits.
Balster and Borzelleca (1982) administered 31.1 mg/kg/day
chloroform in drinking water to pregnant mice throughout gestation and
lactation. No behavioral effects were observed in offspring. Burkhalter
and Balster (1979) administered gavage doses of 31.1 mg/kg/day to adult
ICR mice from 21 days prior to mating until day 21 after birth and to
the pups from days 7 to 21 after birth. Weight gain of pups exposed on
days 7 to 21 after birth was reduced. No significant adverse behavioral
effects were noted.
4.3.3.3 Dermal
Human. No data concerning human developmental effects of dermal
exposure to chloroform were found in the available literature.
Animal. No data concerning animal developmental effects of dermal
exposure to chloroform were found in the available literature.
4.3.3.4 General discussion
Developmental toxicity studies of inhaled and orally administered
chloroform in animals indicate that chloroform is toxic to dams and
fetuses. Chloroform has been shown to cross the placental barrier in
mice. Danielsson et al. (1986) found volatile radioactivity in the
placenta and fetuses of pregnant mice a short time after inhalation of
^C-chloroform. Chloroform metabolites accumulated with time, especially
in the amniotic fluid. In early gestation, metabolites accumulated in
embryonic neural tissues; in late gestation, accumulation occurred in
the fetal respiratory epithelium. Although data regarding developmental
effects of chloroform in humans were not available, there is no reason
to believe that chloroform would not cross the placental barrier in
humans.
4.3.4 Reproductive Toxicity
4.3.4.1 Inhalation
Human. No data concerning human reproductive effects of inhaled
chloroform were found in the available literature.
-------
58 Section 6
Animal. Land et al. (1981) exposed male mice to atmospheric
chloroform concentrations of 0.04 or 0.08% (400 or 800 ppm), 4 h/day for
5 days. Both concentrations caused significant increases in the
percentage of abnormal sperm. Intraperitoneal injection of male mice
with chloroform in corn oil at 0.025 to 0.25 mg/kg/day for 5 days did
not result in increased incidences of sperm head abnormalities (Topham
1980).
4.3.4.2 Oral
Human. No data concerning human reproductive effects of oral
exposure to chloroform were found in the available literature.
Animal. As previously described, Palmer et al. (1979) administered
gavage doses of 0. 15, 30, 150, or 410 mg/kg/day in toothpaste, 6
days/week to SD rats (10/sex/group) for 13 weeks. One of the
histological effects observed in high-dose rats was gonadal atrophy.
4.3.4.3 Dermal
No data concerning human or animal reproductive effects of dermal
exposure to chloroform were found in the available literature.
4.3.4.4 General discussion
There was little useful information concerning the reproductive
effects of chloroform.
4.3.5 Genotoxicity
4.3.5.1 Human
Mixed results were obtained in sister chromatid exchange assays in
cultured human lympohocytes (Table 4.1). No other genotoxicity studies
in humans in vitro or in vivo were found.
4.3.5.2 Nonhuman
Studies on the in vitro genotoxicity of chloroform in prokaryotes,
eukaryotes, and cultured mammalian cells are summarized in Table 4.1.
Results of tests for gene mutation have been negative in prokaryotes and
mixed in yeasts. Chloroform did not produce gene mutations or chromosome
aberrations in hamster cells.
Studies of the in vivo genotoxicity of chloroform are summarized in
Table 4.2. Tests for gene mutations in DrosophLla and for DNA damage in
rats and mice were negative, whereas tests for chromosome aberrations
and sperm abnormalities were mixed.
4.3.5.3 General discussion
The EPA (1985a) summarized studies concerning binding of
metabolically activated chloroform to liver microsomal and nuclear
protein and lipid. They concluded that DNA binding of metabolically
activated chloroform occurs only to a limited degree, if at all.
-------
Table 4.1. Genoloxicity of chloroform in vitro
End point
Gene mutation
DNA damage
Species (test system)
Result
with activation/without activation
Reference
Salmonella typhimurium
Escherichia coli
Yeast
Saccharomyces cerevisiae and
Schizosaccharomyces pombe
Saccharomyces cerevisiae
Chinese hamster lung fibroblasts
Chromosome aberrations Chinese hamster ovary cells
Cytogenetics Human lymphocytes
— /mixed
NT/+fl
NT/-
-/NT
— /mixed
Van Abbe et al. 1982.
Simmon et al. 1977,
Uehleke et al. 1977.
Gockeetal. 1981,
de Serres and Ashby 1981.
and other studies reviewed
by EPA !98Sa and
Rosenthal 1987
Kirkland et al. 1981.
dc Serres and Ashby 1981.
Reitz et al. 1982
Callen et al. 1980
de Serres and Ashby 1981
Reitz et al. 1982
Callen et al. 1980
Sturrock 1977
White et al. 1979
Kirkland et al. 1981.
Morimoto and Koizumi 1983
"NT = not tested.
o
o
o
to
fii
n
to
-------
60 Section 6
Table 4.2. Genotoxicity of chloroform in vivo
End point
Species (test system) Result
References
Gene mutation
Chromosome aberrations
Cytogenetic
Sperm abnormalities
DNA damage
Drosophila melanogaster —
Grasshopper embryos
Mouse bone marrow
Mouse
Rat hepatocytes
Mouse
Gockeet al. 1981,
deSerresand Ashby 1981
+ Liang et al. 1983
Mixed Morimoto and Koizumi 1983,
de Serres and Ashby 1981,
Gockeet al. 1981,
Agustin and Lim-Sylianco
1978
Mixed Topham 1980,
Landetal. 1981,
de Serres and Ashby 1981
— Mirsahset al. 1982
— Reitz et ai. 1980
-------
Toxicological Daca 61
While bacterial mutagenicity tests with chloroform were
predominantly negative, false negative results could have been obtained
because of (1) activation systems inadequate to metabolize chloroform co
a reactive compound, (2) instability and reactivity of phosgene so thac
it may have been scavenged by microsomal protein or lipid before it
could reach the bacterial DNA, and (3) volatility of chloroform that may
have led to inadequate exposure of the bacteria (EPA 1985a; Rosenthal
1987). In addition, a report on the EPA Gene-Tox Program (Kier et al
1986) stated that compounds having complex metabolic pathways, such as
halogenated hydrocarbons, often give false negative results in the
Salmone.Ha/microsome assay. Additional studies are needed before
chloroform could be considered mutagenic in yeast.
Tests for chromosome aberrations, such as sister chromatid
exchange, caused by chloroform in various systems have resulted in mixed
results. Many of these studies are inadequate, and additional studies
may be needed.
The available evidence concerning DNA damage, binding to
macromolecules and mitotic arrest, suggests that chloroform may be
mutagenic, but no definitive conclusion can be reached concerning
mutagenicity of chloroform. The study by Land et al. (1981) reported
significantly increased percentages of abnormal sperm in mice exposed by
inhalation (see Sect. 4.3.4), indicating that chloroform can gain access
to germinal tissue and may pose a mutagenic risk.
4.3.6 Carcinogenicity
4.3.6.1 Inhalation
Data concerning careinogenieity of inhaled chloroform in humans or
animals were not found in the available literature.
4.3.6.2 Oral
Human. Numerous epidemiologic studies of the relationship between
cancer incidence and various components of drinking water, including
chloroform, have been reviewed (EPA 1980, 1985a; NAS 1977). An
association between cancer of the large intestine, rectum, and/or
bladder and the constituents of chlorinated drinking water was evident
in many of these studies. Chloroform has not been identified as the sole
or primary cause of excess cancer rates, but it is one of several
volatile organic contaminants, many of which are considered to have
carcinogenic potential, found in drinking water. Several of the earlier
epidemiologic studies suffered from (1) lack of measured chloroform
concentrations in drinking water; (2) lack of data concerning
concentrations of other organics; (3) limited information concerning
personal drinking water consumption; (4) long latency periods; and (5)
effects of migration, making it difficult to quantify exposure (EPA
198Sa).
Host of the available studies are ecological in nature, but, more
recently, some case-control studies have been published. Most of the
ecological studies tend to support a weak but significant association
between risks of bladder, colon, and rectal cancer and water
chlorination. The odds and risk ratios calculated in the case-control
-------
62 Section 4
studies were as high as 3.6 (Young et al. 1981) but were generally
between 1.1 and 2.0 and could be explained by confounding effects such
as diet, smoking, and occupation. The association appears to be fairly
consistent, however, across several independent and diverse study
groups. Although it can be concluded that the human data suggest a
possible increased risk of cancer at these three sites (from exposure to
chloroform in chlorinated drinking water) because chloroform is the
predominant trihalomethane in drinking water, the data are too weak to
draw a conclusion about the carcinogenic potential of chloroform (EPA
1985a).
Animal. The key carcinogencity studies EPA (1985a) considered for
the derivation of the carcinogenic potency factor are the NCI (1976)
bioassay in rats and mice, a study by Roe et al. (1979) in several
strains of mice, and a study by Jorgenson et al. (1985) in rats and
mice.
NCI (1976) administered gavage doses of chloroform in corn oil to
B6C3F1 mice (50/sex/group) 5 days/week for 78 weeks followed by 14 to 15
weeks of observation. The chloroform used in this study was USP grade,
which is >99.0% chloroform and 0.5 to 1.0% ethanol. Initial dosages were
100 and 200 mg/kg for males and 200 and 400 mg/kg for females. These
levels were increased after 18 weeks to 150 and 300 mg/kg for males and
250 and 500 mg/kg for females. Time-weighted average (TWA) dosages were
138 and 277 mg/kg for males and 238 and 477 mg/kg for females. There was
a statistically significant dose-related increase in hepatocellular
carcinomas in both sexes (females--0/20 controls, 36/45 low dose, and
39/41 high dose, males--1/18 controls, 18/50 low dose, and 44/45 high
dose).
NCI (1976) administered gavage doses of chloroform in corn oil to
Osborne-Mendel rats (50/sex/group) 5 days/week for 78 weeks followed by
a 33-week observation period. Treated male rats received 90 or 180
mg/kg. Treated females initially received 125 and 250 mg/kg, but these
doses were reduced to 90 and 180 mg/kg after 22 weeks and resulted in
TWA doses of 100 and 200 mg/kg administered 5 days/week. There was a
statistically significant dose-related trend in incidence of kidney
epithelial tumors in male rats (0/99 controls, 4/50 low dose, 12/50 high
dose).
Roe et al. (1979) administered gavage doses of chloroform in
toothpaste to four strains of mice (C57BL, CBA, CF/1, and ICI) in three
different experiments. Chloroform was administered 6 days/week for 80
weeks followed by a 13- to 24-week observation period. In one study,
groups of 52 male and 52 female ICI mice received 17 or 60 mg/kg/day in
toothpaste. A group of 100 mice per sex served as controls. In a second
study, groups of 52 male ICI mice received toothpaste containing 0 or 60
rag/kg/day. In the third study, groups of 52 male mice of strains C57CL,
CBA. CF/1, and ICI received 0 or 60 mgAg/day in toothpaste or arachis
oil. In males of the C57BL, CBA, and CF/1 strains, there were no
treatment-related effects on incidence of any type of tumors. There was
a significantly increased incidence of moderate to severe kidney
"changes" in male CBA and CF/1 mice. Male (but not female) ICI mice
experienced an increase in epithelial tumors of the kidney at a dose of
60 mgAg/day (8/38) [but not at 17 mgAg/day (0/37)) relative to
-------
lexicological Data 63
controls (0/72) when chloroform was administered in toothpaste. A more
pronounced increase in kidney tumor incidence was seen in male mice
given 60 mg/kg/day in arachis oil (12/48 treated vs 1/50 vehicle
controls). The incidence of malignant kidney tumors in these mice was
0/50 in controls and 9/48 in the 60 mgAg/da7 group.
Jorgenson et al. (1985) administered 0, 200, 400, 900, and 1800
mg/L chloroform in the drinking water of male Osborne-Mendel rats (50 to
330/group) and female B6C3F1 mice (50 to 430/group) for 104 weeks. There
was also a matched control group of 50 animals whose water intake was
restricted to that of the high-dose group for each species. Group sizes
were adjusted to detect a tumor response at low doses, assuming that
there was a linear relationship between tumor incidence and dose. The
low-dose groups therefore had larger numbers of animals. The chloroform
used in this study was pesticide-quality chloroform that had been
distilled to remove diethylcarbonate. Water consumption of rats and mice
was decreased in a dose-related manner. TWA chloroform dosages reported
by the authors were 0, 19, 38, 81, and 160 mg/kg/day for rats and 0, 34,
65, 130, and 263 mg/kg/day for mice. There were treatment-related
increases in several types of tumors (including renal tumors) in male
rats. However, only the statistically significant increase in incidence
of renal tumors was clearly dose related. Incidences of renal tubular
cell adenomas and/or adenocarcinomas were as follows: 4/301 in controls,
4/313 at 19 mg/kg. 4/148 at 38 mg/kg, 3/48 at 81 mgAg, and 7/50 at 160
mg/kg. The incidence in high-dose male rats was similar to but slightly
lower than that observed in the NCI (1976) corn oil gavage study with
rats of the same strain. Body weight gains were depressed in a dose-
related manner as a result of decreased water and, presumably, food
consumption. Survival was inversely related to chloroform dosage. This
effect was attributed to the increased chloroform dosages being
associated with leaner animals, as indicated by the fact that control
rats on the restricted water intake survived much better than normal
controls.
In mice, a subgroup of treated animals refused to drink water
containing chloroform, resulting in 25% mortality in the two highest -
dose groups during the first 2 weeks of the study. In contrast to the
NCI (1976) mouse study, there was no treatment-related increase in
hepatocellular adenomas and carcinomas in female B6C3F1 mice. No
treatment-related hepatic or renal tumors were found in mice. Jorgenson
et al. (1985) felt that this difference might be due to some interaction
of chloroform with the corn oil vehicle in the NCI (1976) study.
Jorgenson et al. (1985) discussed possible explanations for the
different mouse liver-tumor responses in their study and the NCI (1976)
study. They speculated that either the dosing schedule (bolus gavage
dose vs gradual administration in drinking water) or the vehicle (corn
oil vs drinking water) may have caused the differences. These authors
noted a study by Newbeme et al. (1979) in which corn oil in the diet
increased the number of liver tumors in aflatoxin BI-initiated rats. If
some background level of spontaneously initiated cells is assumed to
occur in B6C3F1 mice, then a chloroform-corn oil interaction might cause
the increased tumor yield. This possibility was supported by Newberne ec
al. (1982), who found that partial hepatectomy increased liver tumors in
B6C3F1 mice. Jorgenson et al. (1985) also noted that the historical
-------
64 Section 4
incidence of liver tumors was consistent with the incidence observed in
their study, and that this incidence had not varied between vehicle
control animals given corn oil by gavage compared with dietary
administration. This indicated that the positive response in the NCI
(1976) study could not be attributed to corn oil alone. Jorgenson et al.
(1985) suggested that corn oil may hasten the expression of chloroform-
initiated cells in the livers of B6C3F1 mice, or chloroform may
interfere with the normal lipid metabolism of the liver.
Pereira et al. (1985) found that chloroform administered in
drinking water appeared to inhibit the development of hepatic tumors Ln
CD-I Swiss mice that had been pretreated with ethylnitrosourea. These
authors speculated that the difference between their results and the NCI
(1976) mouse study may have been caused by the administration of
chloroform as a bolus in the NCI (1976) study or that there may have
been some interaction between chloroform and corn oil in the NCI (1976)
study. Moore et al. (1982) also found that regenerative changes and
toxic effects on liver and kidneys of CFLP outbred Swiss albino mice
were greater when chloroform was administered in corn oil rather than in
a toothpaste base.
Bull et al. (1986) investigated possible vehicle effects on
chloroform toxicity in male and female- B6C3F1 mice. Mice were
administered chloroform by gavage in corn oil or water at doses of 0
60, 130, or 270 mg/kg/day for 91 to 94 days. The chloroform/water
mixtures were 2% emulphor, an emulsifying agent used to produce aqueous
emulsions of hydrophobic chemicals. As detailed in Sect. 4.3.2.1,
chloroform in corn oil was more hepatotoxic than chloroform in emulphor
and more toxic than chloroform administered in drinking water in another
study (e.g., Jorgenson et al. 1985). These results indicated that the
difference in incidence of mouse liver tumors between the NCI (1976) and
Jorgenson et al. (1985) studies was due to the corn oil vehicle, with
chloroform being more hepatotoxic when administered in corn oil than in
water. Because the Withey et al. (1983) study showed that the use of a
corn oil vehicle actually decreased the rate and extent of chloroform
absorption from the gastrointestinal tract. Bull et al. (1986) concluded
that it was unlikely that the differences in carcinogenic response could
be attributed to pharmacokinetic effects. Bull et al. (1986) proposed
that some interaction between the vehicle and chloroform might cause the
difference in carcinogenicity, particularly if the carcinogenic response
in mouse liver is secondary to the hepatoxicity of chloroform.
Data from these and other long-term oral carcinogenicity studies of
chloroform are summarized in Table 4.3.
4.3.6.3 Dermal
No data concerning carcinogenicity of dermal exposure to chloroform
in humans or animals were found in the available literature.
4.3.6.4 General discussion
On the basis of data showing specific carcinogenic activity in
kidneys and liver of mice and rats, EPA (1985a) concluded that there was
sufficient evidence for carcinogenicity of chloroform in experimental
animals using the EPA weight-of-evidence criteria. The EPA (1985a) also
-------
Table 4.3. Oral cucinogcnicily studies of chloroform
Animali
Sprague-Oawley rals
Wislar rais
B6C3FI mice
Strain A mice
Beagle dogs
Osborne-Mendel rals
Exposure
Osborne-Mendel
male rals
0. IS. 75. or 165 mg/kg/day in toothpaste
6 days/week for 52 weeks, or 0 or 60
mg/kg in toothpaste 6 days/week for
80 weeks
2900 ppm in drinking water for 72 weeks.
then I4SO ppm for remainder of lifetime
0. 600. or 1800 ppm in drinking water
for 52 weeks
0. ISO. 300. 600. 1200. or 2400 mg/kg
by gavagc in olive oil once every 4 days
for 30 dose*
0. 15, or 30 mg/kg/day in capsules
6 days/week for 7 years
TWA dosages of 0. 90. and 180 mg/kg/day
(males) and 0. 100. and 200 mg/kg/day
(females) S days/week by gavage in corn
oil for 78 weeks followed by 33 weeks
of observation
0. 200. 400. 900. and 1800 mg/L in
drinking water for 104 weeks
Response
References
No carcinogenic response
Increased incidences of neoplasiic
nodules and hepatic adenofibrosis
compared with control*
No carcinogenic response
Hepalomas in females at 600 and
1200 mg/kg
No carcinogenic response
Increased incidence of kidney
epithelial tumors in males
Palmer el al 1979
Tumasoniset al I98S
Klaumgetal 1986
Eschenbrenner and Miller I945a
Hey wood el al 1979
NCI 1976
Increased incidence of renal tumors Jorgenson el al 198$
5*
o
n>
o
to
r»
ft)
-------
Animals
B6C3FI mice
C57BI, CBA. andCF/l
male mice
ICI mice
NIC mice
Table 4.3 (continued)
in
CD
n
ri
f-.
§
Exposure
TWA dosages of 0. 138. and 271 mg/kg/day
(males) and 0. 138. and 477 mg/kg/day
(females) S days/week by gavagc in corn
oil for 78 weeks followed by 14 to IS weeks
of observation
0. 200. 400. 900. 1800 mg/L in drinking
water for 104 weeks
0 or 60 mg/kg/day 6 days/week by gavage
in toothpaste or arachis oil for 80 weeks
followed by 13 to 24 weeks of observation
0, 17. or 60 mg/kg/day 6 days/week by
gavage in toothpaste or arachis oil for
80 weeks followed by 13 to 24 weeks of
observation
0 I mL of 40% chloroform in oil by
gavage twice weekly for an unspecified
period
Response
Increased incidences of hcpalo-
ccllular carcinomas in males and
females
No increase in liver tumors
No treatment-related effects on
tumor incidence
Increased incidences of epithelial
kidney tumors in males given 60
mg/kg/day in toothpaste or oil
Hepalomas and hepatic lesions in
three of five mice examined
References
NCI 1976
Jorgenson el al 1985
Roe ct al 1979
Roeetal 1979
Rudah 1967
-------
Toxicologies! Data 67
concluded that the epidemlological evidence for chloroform
carcinogenicity was Inadequate. Overall, EPA (1985a) placed chloroform
In Group B2, meaning that it is considered a probable human carcinogen.
IARC (1979, 1982) has classified chloroform in Group 2B.
The recent study by Jorgenson et al. (1985) confirmed the
carcinogenicity of chloroform in producing renal tumors in male
Osborne-Mendel rats, as was observed in the NCI (1976) study. The
primary tumors were of tubular cell origin and occurred at comparable
times in the two studies. Roe et al. (1979) also observed an increase in
renal tumors in ICI mice.
The available data concerning mouse liver tumors are conflicting.
The Jorgenson et al. (1985) drinking water study failed to produce an
increased tumor incidence in livers of female B6C3F1 mice, in contrast
to the NCI (1976) corn oil gavage study. The negative result is
consistent with the absence of liver tumor effects in four strains of
mice tested by Roe et al. (1979), in which chloroform was administered
in a toothpaste base. A pharmacokinetic study by Uithey et al. (1983)
indicated that chloroform was absorbed more slowly and to a lesser
extent from corn oil than from water, suggesting that pharmacokinetic
effects are not responsible for the differences in liver tumor
responses. Historical control data cited by Jorgenson et al. (1985)
indicated that corn oil alone is not responsible for the increased
incidence of liver tumors. Jorgenson et al. (1985) and Bull et al.
(1986) concluded that the corn oil vehicle effect on mouse liver tumors
may be due to some interaction between the vehicle and chloroform.
Another possible explanation for the discrepancy is the difference in
dosing schedule (bolus gavage doses vs gradual dosing in drinking
water). Bolus administration would result in higher-peak blood levels
than gradual drinking water administration of the same dosage. The EPA
(1985a) stated, however, that virtually all of the chloroform
administered in drinking water would be absorbed and metabolized to
reactive metabolites, but a certain percentage of a bolus gavage dose
would be excreted unchanged. Using the available pharmacokinetic data,
EPA (1985a) estimated that for rats and mice given 60 mg/kg by gavage,
the amount of the dose excreted unchanged would be 6% for mice and 20%
for rats. In drinking water administration, virtually all of the dose is
metabolized.
A study by Capel et al. (1979) indicated that chloroform (0.15 or
15 mg/kg/day administered in drinking water for 14 days) could enhance
the growth of three types of murine tumors in mice.
Demi and Oesterle (1985) investigated the promoting activity of
chloroform. Chloroform at doses of 100 to 400 mg/kg administered twice
weekly for 11 weeks caused a dose-related increase in the number of
preneoplastic foci in livers of rats that had been given a single dose
of diethylnitrosamine. Doses of 25 mg/kg had no promoting effect. Doses
of 200 or 400 nig/kg/day for 33 days or 800 mg/kg for 20 days did not
cause formation of preneoplastic foci. These results indicated that
chloroform had promoting rather than initiating activity, although the
initiating potential of chloroform was not adequately assessed in this
-------
68 Section 4
study. A similar study by Pereira et al. (1982) did not detect any
tumor-initiating activity of chloroform, and the results concerning its
promoting activity were inconclusive.
Reltz et al. (1980, 1982) investigated the possible mechanisms of
chloroform carcinogenicity. They found that chloroform doses (single
oral doses of 60 and 240 mg/kg) that resulted in increased tumor
incidences in male B6C3F1 mice also caused severe necrosis at liver and
kidney sites prior to tumor development. A chloroform dose (15 mg/kg)
that did not result in increased tumor incidence also did not cause
necrosis and regeneration. In vivo studies of DNA alkylation and repair
did not indicate any genotoxic effects of chloroform. The authors
suggested that the mechanism of chloroform carcinogenesis was nongenetic
and that noncytotoxic doses should pose no carcinogenic risk.
4.4 INTERACTIONS WITH OTHER CHEMICALS
The EPA (1985a) noted that chloroform toxicity is influenced by
anything that alters microsomal enzyme activity or hepatic GSH levels.
Sato and NakaJima (1984) reviewed data concerning effects of diet
and ethanol on chloroform metabolism and toxicity. They concluded that
food deprivation causes a twofold to threefold increase in hepatic
metabolism of chloroform. The cause of this effect is the lack of
carbohydrate, a high level of which represses the hepatic metabolism of
volatile hydrocarbons. Ethanol was found to increase the hepatic
metabolism of chloroform.
Kutob and Plaa (1962) studied interactions of 1 to 15 daily oral
ethanol doses in mice followed by subcutaneous injections of chloroform
at various times after treatment. They found that the two chemicals
given together caused histological changes in livers that did not occur
with either chemical alone. They also found that ethanol pretreatment
significantly increased the concentration of chloroform in the liver. A
single dose of ethanol was as effective as multiple doses. The authors
proposed that ethanol increased the lipid content of the liver
(increased liver triglyceride content was observed), which results in
increased concentrations of chloroform to be metabolized in the liver. A
similar result was reported by Danni et al. (1981), who found that oral
isopr-panol pretreatment followed by chloroform inhalation caused severe
fatty infiltration of the liver and chloroform alone increased liver
triglycerides.
Sato et al. (1980, 1981) studied the in vitro metabolism of
chloroform by liver microsomal enzyme systems from rats pretreated with
ethanol In drinking water. Their results indicated that ethanol was both
a stimulator and inhibitor of microsomal enzymes, depending on how much
remained in the body and how much time elapsed since ingestion.
Kluwe and Hook (1978) reported that PBBs administered in the diets
of mice potentiated the hepatic and renal toxicity of intraperitoneally
injected chloroform. This effect was assumed to be due to enhanced
chloroform metabolism resulting from PBB Induction of microsomal
enzymes.
-------
lexicological Data 69
Various ketones and ketogenic substances have been found to
potentiate chloroform toxicity. Hewitt et al. (1979) and Cianflone et
al. (1980) found that chlorodecone pretreatment increased hepatotoxicity
of chloroform but that the nonketonic structural analog, mirex, did not
Hewitt et al. (1980) found that other ketones also enhanced hepatic and
renal toxicity of chloroform in rats, with methyl-n-butyl ketone (MBK)
and 2,4-hexandione the most potent, followed by acetone and n-hexane (a
ketogenic chemical) (EPA 1985a). Hewitt et al. (1983) found a positive
correlation between ketone carbon chain length (3C to 7C) and the
severity of potentiated chloroform hepatotoxicity in rats. The ketones
themselves were not hepatotoxic at those doses. Branchflower and Pohl
(1981) proposed that MBK potentiated chloroform hepatotoxicity by
increasing cytochrome P450 levels (thereby enhancing chloroform
metabolism to phosgene) and by decreasing GSH levels.
Disulfiram, an inhibitor of microsomal drug-metabolizing enzymes,
decreases the hepatotoxicity of chloroform (Scholler 1970, Masuda and
Nakayama 1982). Diethyldithiocarbamate and carbon disulfide pretreatment
also protect against chloroform hepatotoxicity (Masuda and Nakayama
1982, 1983; Gopinath and Ford 1975), presumably by inhibiting drug-
metabolizing enzymes.
-------
71
5. MANUFACTURE, IMPORT, USE, AND DISPOSAL
5.1 OVERVIEW
Chloroform is produced at six locations in Che United States
primarily by the chlorination of methyl chloride resulting from the
reaction of methanol and hydrogen chloride. Of the chloroform produced,
93% is used to make fluorocarbon-22. The remainder is either exported or
used as a solvent, fumigant, dry cleaning spot remover, intermediate in
the preparation of dyes and pesticides, and in fire extinguishers.
5.2 PRODUCTION
During 1985, 275.3 million pounds of chloroform were produced
(USITC 1986). Mild summer weather in 1985 caused a decrease in the
demand for air conditioner maintenance and new air conditioners.
Consequently, the output of both fluorocarbon-22 refrigerant and
chloroform was lower in 1985 than in previous years (CMR 1986a).
Manufacturers and sites of chloroform production are as follows:
Occidental Petroleum Corp., Belle, West Virginia; Dow Chemical,
Freeport, Texas, and Plaquemine, Louisiana; LCP Chemicals, Moundsville,
West Virginia; and Vulcan Chemicals, Geismar, Louisiana, and Wichita.
Kansas (SRI 1987). Dupont was scheduled to complete a conversion project
at its Corpus Christi, Texas, plant during 1986 that would add 300
million pounds to its annual chloroform production capacity (CMR 1986a)
Stauffer Chemical is reported to have 72 million pounds per year of idle
capacity at its Louisville, Kentucky, plant (CMR 1986a).
There are two common methods for commercially producing chloroform-
chlorination of methane or chlorination of methyl chloride produced by
the reaction of methanol and hydrogen chloride (Alhstrom and Steele
1979, DeShon 1979). The methanol process accounts for -92% of the
production capacity, and the methane process accounted for only 8% (SRI
1987).
5.3 IMPORT
It appears that chloroform is currently imported into the United
States. Chemical Marketing Reporter (CMR 1986b) indicates that
2,314,851 Ib of chloroform were imported into the United States in
February 1985.
5.4 USES
The use pattern for chloroform is as follows (CMR 1986):
fluorocarbon-22 (F-22), 93% (refrigerants, 52%; fluoropolymers, 41%);
miscellaneous, 4%; and exports, 3%. Miscellaneous uses include use as an
extraction solvent and a solvent for penicillin, alkaloids, vitamins,
-------
72 Section 5
flavors, lacquers, floor polishes, artificial silk manufacture, resins,
fats, greases, gums, waxes, adhesives, oils and rubber, and a dry
cleaning spot remover; in fire extinguishers; as an intermediate in the
preparation of dyes and pesticides; and as fumigants (DeShon 1979).
Chloroform was formerly used as an anesthetic but was replaced by
safer and more versatile materials (DeShon 1979). The U.S. Food and Drug
Administration banned the use of chloroform in drug, cosmetic, and food
packaging products in 1976 (Windholz 1983). This ruling does not include
drug products that contain chloroform in residual amounts from its use
as a processing solvent in manufacture or as a by-product from the
synthesis of drug ingredients (IARC 1979).
5. 5 DISPOSAL
The EPA requires that persons who generate, transport, treat,
store, or dispose of this compound comply with regulations of the
federal Resource Conservation and Recovery Act (RCRA). One method of
disposing of chloroform involves sedimentation followed by dual-media
filtration and adsorption onto activated carbon at 10 to 100 Ib/lb
soluble material (EPA-NIH 1987).
-------
73
6. ENVIRONMENTAL FATE
6.1 OVERVIEW
When released to the atmosphere, chloroform may be transported Long
distances before ultimately being degraded by reaction with
photochemically generated hydroxyl radicals (half-life of -3 months).
Significant amounts of this compound may be removed from the atmosphere
in precipitation; however, most chloroform removed by this mechanism is
likely to reenter the atmosphere by volatilization. When released into
water, volatilization is the primary fate process (half-life of 1-31
days). When released to soil, chloroform will either volatilize rapidly
from the surface or leach readily through soil, ultimately entering
groundwater. It is predicted to persist for relatively long periods of
time in groundwater.
6.2 RELEASES TO THE ENVIRONMENT
As reported by Rehm et al. (1982), the major sources of
anthropogenic release of chloroform to the environment are listed in
Table 6.1. It is assumed that essentially all of the indirectly produced
chloroform would be emitted into the environment. Indirect sources
include bleaching of pulp by pulp and paper mills, drinking water
chlorination, ethylene dichloride manufacture, trichloroethylene
photodegradation, municipal wastewater chlorination, cooling water
chlorination, and automobile exhaust (EPA 1985a,c). Pulp and paper mills
emit more chloroform to the environment than any other single source
Other indirect sources of chloroform that could not be quantified are
chlorination of textile wastewater, the food processing industry,
breweries, combustion of tobacco products treated with chlorinated
pesticides, thermal decomposition of plastics, biological production of
marine algae, and the reaction of chlorinated pollutants with humic
materials in natural waters (EPA 1985a).
6.3 ENVIRONMENTAL FATE
6.3.1 Air
Based on a vapor pressure of 159 mm Hg at 20°C, chloroform is
expected to exist almost entirely in the vapor phase in the atmosphere
(Boublik et al. 1984, Eisenreich et al. 1981). Chloroform has a rather
low partition coefficient between air and water (H - 0.125), which
indicates that significant amounts of this compound may be removed from
the atmosphere in wet deposition (Nicholson et al. 1984). Nevertheless,
most of the chloroform removed from air in precipitation is likely to
reenter the atmosphere by volatilization.
-------
Environmental release (metric
Source
Pulp and paper milli (from the
bleaching of pulp)
Drinking water chlonnalion
Pharmaceutical manufacture
Eihylenc dichlonde manufacture
Tnchloroethylenc photodegradalion
Municipal waitewaler chlonnalion
Cooling water chlonnalion
Mclbyl chloride chlonnalion
Automobile eihauit
Chlorodifluoromethanc manufacture
Loading and traniil losses
Methane chlonnalion
Hypaloo* manufacture
Grain fumigation
Secondary product contamination
Laboratory usage
Total
"Value* are rounded.
Minor releases nossible.
Air
4113
0
570
760
780
0
190
196
180
139
901
70.2
549
28.4
9.8
c
71814
Percent of total*
39
0
5.5
73
75
0
18
19
17
13
09
0.7
OS
0.3
0.1
c
688
Water
298
1.900
46
6
0
320
72
8
0
b
0
33
b
0
06
c
26479
Percent of total*
29
18
04
b
0
31
07
01
0
b
0
003
b
0
0006
c
254
Land
0
0
384
217
0
0
0
54
0
2
0
b
b
0
02
c
6086
tons/year)
Percent of total*
0
0
3.7
21
0
0
0
01
0
002
0
b
b
0
0002
c
58
Total
4.411
1.900
1.000
977
780
320
262
2094
180
141
901
735
549
284
106
c
10.4382"
Percent of total*
42
18
10
94
75
31
25
20
17
14
09
07
05
03
01
c
(0
A
n
it
K-
§
0\
'Not included became of uncertainly
Source Rebm el al. 1982
-------
Environmental Face 75
Reaction of vapor-phase chloroform with photochemically generated
hydroxyl radicals in the -atmosphere is probably the primary degradation
mechanism for this compound in air. The half-life for this reaction in a
typical atmosphere has been estimated to be 70 to 79 days (Atkinson
1985). Chloroform is less reactive in photochemical smog situations (in
the presence of NOX), with a degradation half-life of 260 days
(Diraitriades and Joshi 1977). The relatively slow rate of degradation of
chloroform in air suggests that chloroform vapor has the potential to be
transported over long distances. Assuming a tropospheric-to-
stratospheric turnover time of 30 years, <1% of the tropospheric
chloroform is predicted to diffuse into the stratosphere (Callahan et
al. 1979). Thus, diffusion into the stratosphere is not expected to be
an important fate process.
6.3.2 Water
Chloroform evaporates rapidly from water, with the rate of
evaporation depending on such conditions as rate of reaction,
temperature, and water depth (NLM 1987). Two laboratory studies on the
evaporation of chloroform from stirred (200-2040 rpm) beakers of water
gave volatilization half-lives ranging from 27 min up to 9 h (Rathbun
and Tai 1981; Dilling 1977).
A modeling study of the fate of chloroform in a pond, river,
oligotrophic lake, and eutrophic lake revealed that the dominant removal
mechanism in each case was volatilization. Volatilization half-lives
were predicted to be 36 h, 40 h, 10 days, and 9 days, respectively (EPA
1985a). Based on field monitoring data, the overall half-life of
chloroform has been estimated to be 1.2 days in the Rhine River and 31
days in a lake in the Rhine basin (Zoeteman et al. 1980).
Measured KOc values for chloroform in soil range from 0 to 40,
suggesting that physical adsorption of chloroform to suspended solids
and sediments in water would not be significant (Hutzler et al. 1983) A
modeling study of chloroform in water estimated that the percentage of
total chloroform found in sediments of a typical river, pond,
oligotrophic lake, and eutrophic lake would be 3.07, 8.1, 0.05, and
0.06, respectively (EPA 1985a). This prediction is supported by sediment
monitoring data that indicate that this compound has not been detected
or was detected at low concentrations in sediment (Ferrario et al. 1985,
Helz and Hsu 1978). Field experiments in which chloroform was injected
into an aquifer showed that chloroform was retained poorly by aquifer
material (Roberts et al. 1982).
The hydrolysis half-life of chloroform is >3000 years in water at
pH 7 and 298 K (Mabey and Mill 1978). Since chloroform does not absorb
UV light wavelengths >175 run, this compound is not expected to photolyze
under environmental conditions (A > 290 nm) (Callahan et al. 1979).
Conflicting data are available regarding the biodegradation of
chloroform. Although slow but substantial biodegradation can occur when
the proper microbial populations exist and are acclimated to the
chemical, under aerobic conditions, some studies have shown that little
or no degradation occurs in up to 25 weeks (Bouwer et al. 1981a,
Kawasaki 1980, Heukelekian and Rand 1955). In contrast, other
investigators have reported substantial degradation in much shorter
-------
76 Section 6
periods of time: <49% in 7 days, 100% in 28 days (a large fraction of
this loss was attributed to volatilization), 25% in 14 days, and 67% in
24 days (Tabak et al. 1981, Bouwer et al. 1981b, Flathman and Dahlgran
1982). Under anaerobic conditions, slow degradation has been reported
after acclimation (Bouwer and McCarty 1983). Wilson et al. (1983)
observed no degradation when chloroform was incubated in aquifer
material for 27 weeks.
The bioconcentration factor of chloroform in four different fish
species was found to be <10 times the concentration in ambient water
(Barrows et al. 1980, Anderson and Lusty 1980). This suggests that
chloroform has little or no tendency to bioconcentrate in aquatic
organisms.
6.3.3 Soil
The relatively high vapor pressure of chloroform (159 mm Hg at
20°C), suggests that this compound will volatilize rapidly from dry soil
surfaces (Boublik et al. 1984). Evaporation from moist soil surfaces is
also expected to be significant since this compound does not adsorb to
soil and appears to volatilize fairly rapidly from water.
Chloroform has been found to adsorb strongly to peat moss, less
strongly to clay, very slightly to dolomite limestone, and not at all to
sand (Dilling et al. 1975). The KOc for chloroform in two soils was
measured to be 40, and three other soils with lower organic carbon
content showed no adsorption (Hutzler et al. 1983). The retardation
factor of chloroform applied to a soil column was found to be <1.5
(Wilson et al. 1981). These data suggest that chloroform would be highly
mobile in soil.
Based on the data in aquatic media, the chemical reaction of
chloroform in soil does not appear to be a significant fate process.
No data on the biodegradation of chloroform in soil were found in
the available literature. Based on data in aquatic media, it appears
that chloroform may biodegrade to some extent under both aerobic and
anaerobic conditions, provided that suitable microbial po-ilations are
present and that acclimitization to the chemical has occurred.
-------
7. POTENTIAL FOR HUMAN EXPOSURE
7.1 OVERVIEW
Chloroform is both a man-made and naturally occurring compound,
although anthropogenic sources are responsible for most of the
chloroform found in the environment. Most of the chloroform released co
the environment will eventually end up in the atmosphere, whereas much
smaller amounts will eventually end up in groundwater. In the
atmosphere, chloroform may be transported long distances before
ultimately being degraded by photochemical reaction. This has been
substantiated by the detection of chloroform in ambient air in remote
locations far removed from anthropogenic sources. Chloroform leaches
into groundwater primarily from spills, landfills, and industrial
sources. Upon contamination of groundwater, chloroform is expected to
persist for relatively long periods of time.
The general population is exposed to chloroform by ingestion of
drinking water, inhalation, and consumption of many foods. The average
daily intake of chloroform from both ingestion of finished drinking
water and inhalation has been estimated to range from 64 to 396 /ig/day
(see Sect. 7.2.1 and 7.2.2 on levels monitored or estimated in the
environment in air and in water). Diet may also contribute signif icar>cl
to the daily intake of chloroform.
A 1981-83 NIOSH survey estimated that 87,600 workers are
potentially exposed to chloroform in the United States; however, this
figure does not include exposure to trade-name products that contain
this compound. Occupational exposure is expected to occur primarily by
the inhalation and dermal routes.
7.2 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT
7.2.1 Air
Typical U.S. environmental background levels of chloroform in
outdoor air in rural/remote, urban/suburban, and source-dominated areas
are 0.02 to 0.2, 0.2 to 3.4 and 0.2 to 13 Mg/m3, respectively
(Brodzinsky and Singh 1982, Class and Ballschmidter 1986, Bozzelli and
Kebbekus 1979, Singh et al. 1982, Harkov et al. 1984, Wallace 1986)
Typical concentrations in indoor air were found to range between 0.07 co
3.6 Mg/n3 (Wallace et al. 1986a, Andelman 198Sa,b). A few investigators
reported the levels of chloroform in personal inhaled air (both indoor
and outdoor air) and exhaled air (Wallace et al. 1986a,b).
Concentrations well above typical background levels have been found. For
example, 110 Mg/n>3 was detected in air samples collected outside of
homes in "Old Love Canal" in Niagara Falls, New York (Barkley et al.
-------
78 Section 7
1980). Chloroform has also been detected in rainwater at concentrations
as high as 0.25 pg/L (Kawamura and Kaplan 1983).
Assuming that the average daily intake of air is 20 m3/day typica'
values for the average daily intake of chloroform by inhalation'in
urban, rural, and source-dominated areas have been estimated to be 0 4
to 4.0, 4.0 to 68. and 4.0 to 260 pg, respectively.
7.2.2 Water
Chloroform has been monitored at widely varying concentrations in
surface water, groundwater, and drinking water throughout the United
States. Concentrations as high as 0.37 Mg/L in surface water. 490 ug/L
in groundwater. and 21,800 Mg/L in landfill leachate have been measured
(Strachan and Edwards 1984, Rao et al. 1985, DeWalle and Chian 1981)
Chloroform has also been found in a limited number of sediment samples
with a maximum concentration of 18 MgAg (w/w) detected in sediments
from Lake Pontchartrain in New Orleans, Louisiana (Ferrario et al.
1985) .
The level of chloroform in drinking water in the United States was
determined during the 1975 EPA National Organics Reconnaissance Survey
(NORS) and the 1976-77 EPA National Organic Monitoring Study (NOUS)
Drinking water samples from a total of 137 cities geographically
distributed across the United States were studied. Combined results of
the NORS and NOMS reveal that chloroform was detected in 99.5% of the
finished drinking water samples. Concentrations ranged from below
detection limits to 311 pg/L. with concentrations in most of the samples
ranging between 32 and 68 Mg/L (Brass et al. 1977, Symons et al. 1975)
Assuming that the average person consumes 2 L of water per day, the
average daily intake of chloroform by ingestion of water has been
estimated to be 64 to 136 /ig.
The main source of chloroform found in municipal drinking water is
the chlorination of naturally occurring humic materials found in raw
water supplies (Cech et al. 1982. Bellar et al. 1974). The concentration
of chloroform in drinking water has been found to increase with time
(Kasso and Wells 1981). Thus, the concentration of chloroform increases
as water moves through distribution systems (Varma et al. 1984).
7.2.3 Soil
Monitoring data for chloroform in soil were not located in the
available literature.
7.2.4 Other
Chloroform has been detected in various foods at the following
levels: seafood, 3 to 180 MgAg; dairy products, 1.4 to 56 ^gAg: meat
1 to 4 MgAg; oil/fats, 2 to 24 pgAg; beverages, 0.4 to 178 pgAg;
fruits/vegetable, 2 to 18 ,ig/kg; bread. 2 jigAg; and mother's milk, not
quantified (EPA 1980, McConnell et al. 1975, Entz et al. 1982, Lovegren
et al. 1979, Coleman et al. 1981, Pellizzarl et al. 1982). Although it
appears that the dietary intake of chloroform may be substantial, data
are insufficient to predict the daily average.
-------
Potential for Human Exposure 79
In April 1976, the U.S. Food and Drug Administration (FDA)
identified approximately 1900 drug products for humans that contained
chloroform. Such products included toothpastes, cough syrups,
expectorants, antihistamines, liniments, and decongestants. In July
1976, the FDA banned the use of chloroform as an ingredient (active or
inactive) in drug and cosmetic products. However, chloroform is not
considered to be an ingredient if drug products contain residual amouncs
of the compound as a result of its use as a processing solvent in
manufacture or its formation as a by-product from the synthesis of
another ingredient (IARC 1979). This suggests that many commodities,
which the general population may come in frequent contact with, may
contain residual amounts of chloroform.
Chloroform is typically found in automobile exhaust at a
concentration of 0.027 mg/m3 (EPA 1980). Chloroform has been detected in
the air above outdoor and indoor pools and in spas, with average
concentrations ranging between <0.1 to 68, 0.5 to 274, and <0.1 to
130 /ig/m3, respectively. Average chloroform levels in water samples
taken from outdoor pools, indoor pools, and spas were 103 to 158, 8 to
350, and 2 to 292 /*g/L, respectively (Armstrong and Golden 1986). The
use of chlorine-containing bleaches and/or scouring powders for washing
clothes and dishes or during other household activities may produce
elevated levels of chloroform. Similarly, the indoor use of certain
rodenticides may produce higher levels of chloroform in indoor air
(Wallace et al. 1987). Volatilization of chloroform from water during
showering may be a substantial source of inhalation exposure (Andelman
1985a,b; Wallace, 1986).
7.3 OCCUPATIONAL EXPOSURES
A National Occupational Hazard Survey (NOHS) conducted between 19"2
and 1974 estimated that 215,000 workers in the United States are
potentially exposed to chloroform (NIOSH 1984). This figure includes 46-
of observations termed as "generic" by NIOSH, that is, the surveyor
observed a product in some type of general use that led NIOSH to suspecc
that the specific agent (in this case chloroform) was contained in thac
product. A survey conducted between 1981 and 1983 estimated that 86,700
workers in the United States are potentially exposed to chloroform
(NIOSH 1987b). This figure is based on actual observations only and does
not include exposure to trade-name products that contain Chloroform.
Results of a study of 350 subjects living in Bayonne, New Jersey,
indicate that the breath levels and personal air exposures of people who
currently work in a paint store, had recently been in a paint store, or
had recently used pesticides were significantly elevated in comparison
to people not involved in these activities (Wallace 1986). Based on the
physical properties and uses of chloroform, it has been concluded thac
occupational exposure would occur primarily by inhalation and dermal
contact.
7.4 POPULATIONS AT HIGH RISK
Since chloroform is a component of chlorinated drinking water,
people whose water supplies are chlorinated have the potential for
exposure. Raw surface water supplies are usually disinfected by
chlorination, although raw groundwater supplies (from deep wells) are
-------
80 Section 7
commonly distributed with no disinfection (Singley 1984). Approximately
12,000 of 60,000 public water supplies in the United States use surface
water as a source for raw water. These systems supply 66% of the
population using public water supplies (EPA 1985d).
As reviewed by EPA (1985a), several factors, including ethanol
ingestion and starvation, can potentiate the toxicity of chloroform to
the liver. Therefore, people who drink alcohol or who diet may be at
higher risk.
-------
8. ANALYTICAL METHODS
Several methods are available for the analysis of chloroform in
different environmental and biological matrices The choice of a
particular method will depend on the nature of the sample matrix, the
required precision, accuracy, and detection limit, and the cost and
turnaround time of the analysis. Biological samples for a wide spectrum
of halogenated volatile substances, including chloroform, can be
screened by injecting static head-space gas or solvent-extracted liquid
into a gas chromatograph equipped with an electron capture detector
(Foerster and Garriott 1981, NYDEC 1985). Preconcentrations of samples
prior to quantification increase the detection limits of the method
used. In air samples, the preconcentration usually is done by passing
the air through a suitable adsorbent during sample collection. The
dynamic headspace technique commonly known as purge-and-trap provides an
excellent method for the preconcentration of chloroform from water,
food, soil, sediment, and biological matrices. The purge-and-trap method
also provides a preliminary separation of chloroform from other less
volatile and nonvolatile components in the samples, thereby eliminating
the need for extensive separation of the components by a gas
chromatographic column prior to quantification. The best specificity a-^c
sensitivity for chloroform quantification is obtained by an electrolyc :..-
conductivity detector in the halide detection mode, since this deteccor
is relatively insensitive to nonhalogenated species and very sensitive
to halogenated species. Although mass spectrometry is less sensitive
than an electrolytic conductivity detector, it is often used as a
confirmation method, since the ion-chromatograms of the fragment and
parent ions provide an alternative confirmation in addition to the gas
chromatographic retention time. For details of the analytical methods,
including the advantages and disadvantages, specificity,
reproducibility, and sensitivity, refer to the references cited in
Table 8.1.
8.L ENVIRONMENTAL MEDIA
Analytical methods and detection limits for chloroform in various
environmental matrices are identified in Table 8.1. These methods are
based primarily on gas chromatographic separation, with subsequent
quantification by various methods. Table 8.1 includes the EPA (1987a)
methods required by the EPA Contract Laboratory Program (CLP) for
analysis of chloroform in water, soil, and sediment.
8.2 BIOMEDICAL SAMPLES
Analytical methods and detection limits for chloroform in various
biomedical samples are identified in Table 8.1. Biological samples are
commonly pretreated for analysis by a purge-and-trap method or a
-------
Table 8.1. Analytical methods for cbloroTona
Sample matrix
Air
Ambicni air and slacks
Air
Ekhaled air (breath)
Groundwater, liquid and
solid matrices
Wasicwiter
Blood, urine, and tissue
Volatile food components
Drinking water
Sample preparation"
Adsorption on charcoal, desorp-
two with carbon disulfide
Tcoai OC adsorption and thermal
doorption
None
Sample collected in Tedlar bag
preconcentrated by Tenax-GC.
thermal doorptioo
Direct injection of headspace
gas (EPA method 5020) or
preconceniraiion by purge-and-
trap and thermal desorption
(EPA method 5030)
Preconceniraiion by purge-and-
trap method and thermal desorp-
non
Purgc-and-irap, thermal dcsorp-
tion
Direct injection of headspace
gas
Direct injection or purge-and-
trap on GC column (automated)
Solvent extraction
Purge-and trap and thermal
dcsurplion
Quantification method"
GC/FID
(NIOSH method 1003)
GC/FID or ECD
GC/ECD
HRGC/FID and HRGC/MS
GC/HSD or FID
(EPA method 8010)
GC/HSD or MS
(EPA methods 601
and 624)
GC/HSD or GC/MS
GC/ECD or MS
GC/ECD.
GC/Hall
GC/HD.
GC/MS
dC/MS
Detection limit
0 7 mg/m1
for 15-L sample
NR*
1 5 Mg/m'
NR
OOSjig/L
005Mg/L(forHSD)
1 6 Mg/L (for MS)
010 Mg/L
(blood and urine)
4 2 Mg/kg
(beverages)
12 5 Mg/kg
(dairy products)
18 «.g/kg
(meals)
28 jig/kg
(fats/oils)
1 «ig/L (direct)
0 1 Mg/L (purge-
and-trap)
004/ig/L
NR
0 1 ^g/L
Accuracy/Recovery
97% ai
100-500 mg/m1
NR
NR
NR
102% al 044-50 ^g/L
102% alO 44-50 ^g/L
(for HSD)
101% at 10-100 pg/L
(for MS)
NR
NR
NR
NR
NR
103- 126% at 3S-70
Mg/L (direct)
91-106% at 35-70
-------
Tible8.l (continued)
Sample maim
Drinking water
Scawaicr and freshwater
Water
Water
Blood
Whole blood
Adipose tissue and serum
Tap water and whole blood
Serum and adipose tissue
Water, serum, and urine
Water, soil, and sediment
Water, soil, and sediment
fl
Sample preparation0
None (direct injection)
None (direct injection)
Solvent extraction with
pcntane
Permeation through a silicon
polycarbonate membrane
Solvent extraction
headspace analysis
Purge-and-lrap and thermal
desorplion
Purge-and-lrap and thermal
dcsorption
Solvent extraction
Purge-and-trap and thermal
desorplion
Solvent extraction
Solvent extraction
r
Purge-and-lrap and thermal
desorplion
Quantification method"
GC/ECD
GC/MS
GC/ECD
GC/FID
GC/ECD
GC/FID
GC/MS
GC/Hall with GC/MS
confirmation
HRGC/ECD with HRGC/MS
confirmation
GC/Hall with GC/MS
confirmation
HRGC/ECD
GC/hID (screening lest)
(EPA-CLP)
GC/MS
(EPA-CLP)
Detection limit
5
to
i— .
n
o
to
K.
a
n
n
O
Q
in
"GC. gas chromatography. HRGC, high-resolution gas chromaiography, FID. flame lomzaiion detector. I ISP, electrolytic conductivity detector, ECD. electron capture dctec-
lor. Hall. Hall conductivity deteclor. MS, mass spectrometry
6 Not reported
-------
84 Section 8
variation of that method (EPA 1985a). Details of the pretreatment and
quantification methods are provided in the cited references.
Caution should be exercised in interpreting the results from the
analysis of chloroform levels in body tissues. Tetrachloroethylene,
trichloroethylene, trichloroacetaldehyde, and trichloroacetic acid can
all be metabolic precursors of chloroform (Peoples et al. 1979). The
presence of any of these compounds in body tissues may cause
artifactually higher values for chloroform. Heating of tissues during
treatment of samples as a part of the analytical procedure may produce
artifactually higher values of serum chloroform in the presence of
serum-bound trichloroacetic acid because of thermal degradation of the
latter compound to chloroform. It is probably for these reasons that a
poor correlation between chloroform exposure and tissue levels has been
found.
-------
85
9. REGULATORY AND ADVISORY STATUS
9 . 1 INTERNATIONAL
IRPTC (1987) reported a World Health Organization (WHO) drinking
water guideline of 30 pg/L for chloroform.
9 . 2 NATIONAL
9.2.1 Regulations
The current OSHA PEL for chloroform is 50 ppm in the workroom
atmosphere (OSHA 1985) . The EPA amended the National Interim Primary
Drinking Water Regulations, adding a section on the control of organic
halogenated contaminants. The limit for total trihalomethanes , including
chloroform, was 0.1 mg/L. This limit was based partly on estimates of
cancer risk and partly on technical and economic feasibility. It applies
only to water supplies serving more than 10,000 consumers (EPA 1980
1985a).
The EPA (1980) derived cancer-based ambient water quality criteria
for chloroform. A potency estimate from the NCI (1976) bioassay with
female mice was used to derive these criteria. Chloroform levels
associated with incremental increases of cancer risk of 10*5, 10'6, and
10'7 were 1.9, 0.19, and 0.019 /ig/L, respectively, if exposure is
assumed to be from drinking water and from consuming fish and shellfish
from contaminated ambient water. If exposure is assumed to be from
consumption of contaminated fish and shellfish only, the corresponding
criteria are 157, 15.7, and 1.57 Mg/L in ambient water.
Chloroform levels in drinking water associated with incremental
increases of cancer risk of 10"4, 10'5, 10'6, and 10'7 are 600, 60, 6,
and 0.6 ^g/L, respectively (EPA 1987b) .
Chloroform is regulated by the Clean Water Act Effluent Guidelines
for the following industrial point sources: electroplating, organic
chemicals, steam electric, asbestos, timber products processing, metal
finishing, paving and roofing, paint formulating, ink formulating, gum
and wood, carbon black, metal molding and casting, coil coating, copper
forming, and electrical and electronic components (EPA 1988).
Federal law (CERCLA 103a and 103b) requires that the National
Response Center be notified when there is a release of a hazardous
substance in excess of the reportable quantity (RQ) . The RQ for
chloroform is 5000 Ib (EPA 1985e) ; however, this RQ is subject to change
when the carcinogenicity and/or toxicity assessment is completed (EPA
1987c).
-------
86 Section 9
Federal law (Section 302 of SARA) requires any facility where an
extremely hazardous substance is present in excess of the threshold
planning quantity (TPQ) to notify the state emergency planning
commission. The TPQ for chloroform is 10.000 Ib (EPA 1987c). Federal la-
Section 304 of SARA) also requires immediate reporting of releases of
hazardous substances to local emergency planning committees and the
state emergency planning commission.
Federal law (Section 313 of SARA) requires owners and operators of
certain facilities that manufacture, process, or otherwise use
chloroform to report annually to both EPA and the state in which the
facility is located their releases of chloroform to the environment (EPA
1987d).
The EPA intends to list chloroform as a hazardous air pollutant
under Section 112 of the Clean Air Act (EPA 1985f).
The FDA approved chloroform for use as an indirect food additive
(i.e., a component of articles that may come in contact with food.) It
has been exempted from tolerance when used as a solvent in pesticide
formulations applied to crops. The FDA has restricted its use in drugs
and cosmetic products as a result of the positive NCI (1976) bioassay
(EPA 1985a).
9.2.2 Advisory Guidance
9.2.2.1 Air
NIOSH (1974) recommended that atmospheric chloroform concentrations
not exceed 10 ppm as a TWA for up to a 10-h workday, 40-h workweek.
NIOSH (1974) also proposed a 10-min ceiling level of 50 ppm. These
criteria were designed to protect against mild CNS depression,
irritation, and fetal abnormalities (which were considered to occur at
concentrations lower than those causing liver damage). The NIOSH
criterion was lowered to 2 ppm in 1976 in response to the positive
results of the NCI (1976) bioassay (NIOSH 1977). This level applied to
halogenated anesthetics, including chloroform, and was selected because
it was the lowest level detectable with sampling and analysis
techniques, and not because any safe level of chloroform exposure had
been defined (EPA 198Sa).
ACGIH (1986) recommended a TLV-TWA of 10 ppm to protect against
carcinogenicity and embryotoxicity as a result of occupational exposure
to chloroform in the atmosphere. ACGIH (1986) gave chloroform an A2
classification--substances suspect of carcinogenic potential for man.
Chloroform levels in air associated with incremental increases of
cancer risk of 10'4, 10'5 10'6 and 10'7 are 4.3 x 10'3, 4.3 x 10'4.
4.3 x 10'5, and 4.3 x 10'6 mg/m3 (8.8 x 10'4. 8.8 x 10'5, 8.8 x 10'6.
and 8.8 x 10'7 ppm), respectively (EPA 1985a).
-------
Regulacory and Advisory Status 37
9.2.3 Data Analysis
9.2.3.1 Reference dose
The EPA (1987d) has calculated a reference dose (RfD) of 0 01
mg/kg/day for chloroform. This value is based on the LOAEL of 12.9
mg/kg/day for increased fatty cyst formation in livers of dogs exposed
to 15 mg/kg/day, 6 days/week, for 7.5 years (Heywood et al. 1979). The
RfD is calculated according to the methods of Barnes et al. (1987) as
follows:
RfD - 15 mg/kg/day x 6 days/7 days/(100) (10) - 0.01 mg/kg/day
where: 15 mg/kg/day - LOAEL
6 days/7 days - factor to expand over a 7-day week
100 - uncertainty factor for inter- and
intraspecies extrapolation
10 - uncertainty factor appropriate for using
a LOAEL in the absence of a NOAEL
9.2.3.2 Carcinogenic potency
The EPA (1985a) performed a quantitative carcinogenicity risk
assessment for chloroform. Five data sets were analyzed: incidences of
liver tumors in female mice (NCI 1976), liver tumors in male mice (NCI
1976), kidney tumors in male rats (NCI 1976), kidney tumors in male mice
(Roe et al. 1979), and kidney tumors in rats (Jorgenson et al. 1985).
For inhalation exposure. EPA (1985a) chose to combine the potency
estimates from NCI (1976) liver tumor data for male and female mice The
geometric mean of these two estimates resulted in the q * of 8.1 x 10"2
(mg/kg/day)'1. Using this q * EPA (1985a) calculated upper-bound
estimates of cancer risk for exposure to 1 Mg/m^ in air to be 2.3 x
10'5. This q.* for inhalation exposure to chloroform was validated by
the CRAVE (Carcinogen Risk Assessment Verification Endeavor) work group
on August 26, 1987 (EPA 1987b). The CRAVE work group (EPA 1987b) also
validated a q * for oral exposure via drinking water of 6.1 x 10*3
(mg/kg/day)"^ based on the incidence of kidney tumors in male rats in
the study by Jorgenson et al. (1985). The upper-bound estimate of cancer
risk for exposure to 1 pg/L in water is 1.7 x 10~7.
Chloroform has been classified as a Group B2 carcinogen, that is, a
probable human carcinogen, based on sufficient evidence from animal
studies and inadequate evidence from human studies (EPA 1986a, 1987b).
IARC (1979, 1982) classified chloroform in Group 2B.
9.3 STATE
Regulations and advisory guidance from the states were not
available.
-------
89
10. REFERENCES
ACGIH (American Conference of Governmental Industrial Hygienists). 1986.
Documentation of the Threshold Limit Values and Biological Exposure
Indices. 5th ed. Cincinnati, OH, p. 130.
Adriani J. 1970. The pharmacology of anesthetic drugs. Thomas CC, ed.
Springfield, IL: pp. 57-60 (cited in EPA 1985a).
Agustin JS, Lim-Sylianco CY. 1978. Mutagenic and clastogenic effects of
chloroform. Bull Phil Biochem Soc 1:17-23 (cited in EPA 1985a).
Ahlstrom RC, Steele JM. 1979. Chlorocarbons, hydrocarbons (CH3C1). In:
Grayson M, Eckroth D, eds. Kirk-Othmer Encyclopedia of Chemical
Technology, 3rd ed. Vol 5. New York: John Wiley and Sons, pp. 677-685.
Ahmed AE, Kubic VL, Anders MU. 1977.. Metabolism of haloforms to carbon
monoxide. I. In vitro studies. Drug Metab Dispos 5:198-204.
Amoore JE, Hautala E. 1983. Odor as an aid to chemical safety: Odor
thresholds compared with threshold limit values and volatilities for 214
industrial chemicals in air and water dilution. J Appl Toxicol
3:272-290.
Andelman J. 1985a. Human exposures to volatile halogenated organic
chemicals in indoor and outdoor air. Environ Health Perspect 62:313-318
Andelman J. 1985b. Inhalation exposure in the home to volatile organic
contaminants of drinking water. Sci Total Environ 47:443-460.
Anders MW, Stevens JL, Sprague RV, Shaath Z. 1978. Metabolism of
haloforms to carbon monoxide. II. In vivo studies. Drug Metab Dispos
6:556-560.
Anderson DR, Lusty EB. 1980. Acute Toxicity and Bioaccumulation of
Chloroform to Four Species of Freshwater Fish: Sal/no gairdneri, Rainbow
Trout, Lepoals macrochirus, Bluegill....Report CR-0893. Nuclear
Regulatory Commission, NTIS PNL-3046 (cited in NLM 1987).
*Key studies.
-------
90 Section 10
Antoine SR, DeLeon IR, O'Dell-Smith RM. 1986. Environmentally
significant volatile organic pollutants in human blood. Bull Environ
Contain Toxlcol 36:364-371.
Armstrong DW, Golden T. 1986. Determination of distribution and
concentration of trihalomethanes in aquatic recreational and therapeutic
facilities by electron capture GC. LC-GC 4(7):652-655.
Atkinson R. 1985. Kinetics and mechanisms of the gas-phase reactions of
nydroxyl radical with organic compounds under atmospheric conditions
Chem Rev 85:69-201.
Bailie MB, Smith JF, Newton JH, Hook JB. 1984. Mechanism of chloroform
nephrotoxicity. IV. Phenobarbital potentiation of in vivo chloroform
metabolism and toxiclty in rabbit kidneys. Toxicol Appl Pharmacol
74:285-292 (cited in EPA 1985a).
Balster RL, Borzelleca JF. 1982. Behavioral toxicity of trihalomethane
contaminants of drinking water in mice. Environ Health Perspect
^€> 112 7 * 13o.
Barkley J et al. 1980. Gas chromatography mass spectrometry computer
analysis of volatile halogenated hydrocarbons in man and his
environment. A multimedia environmental study. Biomed Mass Spectrom
/ • Ai J * m il / •
Barnes D et al. 1987. Reference dose (RfD): Description and use in
health risk assessments. Appendix A of the Integrated Risk Information
System (IRIS). Washington, DC: Office of Health and Environmental
Assessment, Office of Research and Development, EPA 600/8-86-0321.
Barrows ME, Petrocelli SR, Macek KJ, Carroll JJ. 1980. Bioconcentration
and elimination of selected water pollutants by bluegill sunfish
(Lepomis macrochirus). In Dyn, Exposure Hazard Assess Toxic Chem. Ann
Arbor, MI: Ann Arbor Science, pp. 379-392.
Bellar TA, Llchtenberg JJ, Kroner RC. 1974. The occurrence of
organohalides In chlorinated drinking water. J Am Water Works Assoc
66:703-706 (cited In EPA 1985a).
Bennet RA, Whlgham A. 1964. Chloroform sensitivity of mice. Nature
204:1328.
Blanchard RD. Hardy JK. 1986. Continuous monitoring device for the
58(7V1529 15 " V°latlle or«anlc Priority pollutants. Anal Chem
* Bomskl H, Sobolweska A, Strakowskl A. 1967. Toxic damage of the liver
J? ?5;0f°form ln chemlcal Industry workers. Arch Gewerbepathol Gewerbehy
24:127-134 (FRG) (translated from German).
-------
References 9 "..
Boublik T, Fried V, Hala E. 1984. The vapor pressures of pure
substances: Selected values of the temperature dependence of the vapor
pressures of some pure substances in the normal and low-pressure region
Vol. 17. Amsterdam, Netherlands: Elsevier Sci Publ.
Bouwer EJ, McCarty !*L. 1983. Transformations of 1- and 2-carbon
halogenated aliphatic organic compounds under methanogenic conditions
Appl Environ Microbiol 45(4):1286-1294.
Bouwer EJ, McCarty PL, Lance JC. 1981a. Trace organic behavior in soil
columns during rapid infiltration of secondary wastewater. Water Res
15:151-159.
Bouwer EJ, Rittman B, McCarty PL. 1981b. Anaerobic degradation of
halogenated 1- and 2-carbon organic compounds. Environ Sci Technol
15:596-599.
Bowman FJ, Borzelleca J, Munson AE. 1978. The toxicity of some
halomethanes in mice. Toxicol Appl Pharmacol 44:213-215.
Bozzelli JW, Kebbekus BB. 1979. Analysis of selected volatile organic
substances in ambient air, final report Apr-Nov. 1978. Newark, NJ: New
Jersey Institute of Technology.
Branchflower RV, Pohl LR. 1981. Investigation of the mechanism of the
potentiation of chloroform-induced hepatotoxicity and nephrotoxicity by
methyl n-butyl ketone. Toxicol Appl Pharmacol 61:407-413 (cited in EPA
1985a).
Branchflower RV, Nunn DS, Highet RJ, Smith JH, Hook JB, Pohl LR. 1984
Nephrotoxicity of chloroform: Metabolism to phosgene by the mouse
kidney. Toxicol Appl Pharmacol 72:159-168.
Brass HJ, Feige MA, Halloran T, Mello JW, Munch D, Thomas RF. 1977. The
National Organic Monitoring Survey: Sampling and analyses for purgeable
organic compounds. In: Drinking Water Quality Enhancement Source
Protection, pp. 393-416.
Brodzinsky R, Singh HB. 1982. Volatile organic chemicals in the
atmosphere: An assessment of available data. Menlo Park, CA: Atmospheric
Science Center, SRI International. Contract 68-02-3452.
Brown DM, Langley PF, Smith D, Taylor DC. 1974a. Metabolism of
chloroform. I. The metabolism of l^C-chloroform by different species
XenobioCica 4:151-163.
Brown BR Jr, Sipes IG, Sagalyn AM. 1974b. Mechanisms of acute hepatic
toxicity. Chloroform, halothane, and glutathione. Anesthesiology
41:554-561.
Bull RJ et al. 1986. Enhancement of the hepatotoxicity of chloroform in
B6C3F1 mice by corn oil. Implications for chloroform carcinogenesis.
Environ Health Perspect 69:49-58.
-------
92 Section 10
Bureau International Technique des Solvents Chlores. 1976.
Standardization of methods for the determination of traces of some
volatile chlorinated aliphatic hydrocarbons in air and water by gas
chromatography. Anal Chim Acta 82:1-17 (cited in IARC 1979).
Burkhalter J, Balster RL. 1979. Behavioral teratology evaluation of
chloroform in mice. Neurobehav Toxicol 1:199-205 (cited in EPA 1985a)
Caldwell KK, Harris RA. 1985. Effects of anesthetic and anticonvulsant
drugs on calcium-dependent efflux of potassium from human erythrocytes
Eur J Pharmacol 107(2) : 119-125.
Callen DF, Wolf CR, Philpot RM. 1980. Cytochrome P-450 mediated genetic
activity and cytotoxicity of seven halogenated aliphatic hydrocarbons in
Saccharomyces cerevLsiae. Mutat Res 77:55-63.
Callahan MA, Slimak MW, Gabel NW, et al. 1979. Water-related
environmental fate of 129 priority pollutants. Vol. II. Washington DC-
EPA. EPA-440/4-79-029B.
Capel ID, Dorrell HM, Jenner M. Pinnock MH, Williams DC. 1979. The
effect of chloroform ingestion on the growth of some murine tumors Eur
J Cancer 15:1485-1490.
Cech I, Smith V, Henry J. 1982. Spatial and seasonal variations in
concentration of trihalooethanes in drinking water. In: Albaiges J, ed.
Analytical Techniques in Environmental Chemistry. II. New York- Pereamon
Press, pp. 19-38.
* Challen PJR, Hickish DE, Bedford J. 1958. Chronic chloroform
intoxication. Br J Ind Med 15:243-249.
Chemline. 1987. On-line Computer Data Base. National Library of
Medicine. Retrieval Data 6/87.
Chenoweth MB., Robertson DN, Erly DS, Goekhe K. 1962. Blood and tissue
levels of ether, chloroform, halothane, and methoxyfluorane in dogs
Anesthesiology 23:101-106.
Chiou WL. 1975. Quantitation of hepatic and pulmonary first-pass effects
and its implication in pharmacokinetic study. I. Pharmacokinetics of
chloroform in man. J Pharmacokinet Biopharm 3:193-201.
Chu I, Secours V, Marino I, Villeneuve DC. 1980. The acute toxicity of
four trihalomethanes in male and female rats. Toxicol Appl Pharmacol
52:351-353.
* Chu I, Villeneuve DC, Secours VE, Becking GC. 1982a. Toxicity of
trihalomethanes. I. The acute and subacute toxicity of chloroform,
bromodichloromethane, chlorodibromomethane, and bromoform in rats
J Environ Sci Health B17:205-224.
-------
References 93
* Chu I, Villeneuve DC, Secours VE, Becking GG, Valli VE. 1982b.
Toxicity of trihalomethanes. II. Reversibility of toxicological changes
produced by chloroform, bromodichloromethane, chlorodibromomethane, and
bromoform in rats. J Environ Sci Health 817:225-240.
Cianflone DJ, Hewitt WR. Villeneuve DC, Plaa GL. 1980. Role of
biotransformation in the alterations of chloroform hepatotoxicity
produced by Kepone and Mirex. Toxicol Appl Pharmacol 53:140-149 (cited
in EPA 1985a).
Class T, Ballschmidter K. 1986. Chemistry of organic traces in air. VI
Distribution of chlorinated Cl-C4-hydrocarbons in air over the northern
and southern Atlantic Ocean. Chemosphere 15(4):413-427.
CMR (Chemical Marketing Reporter). 1986a. Chemical Profile: Chloroform
New York: Schnell Publishing, February 17.
CMR (Chemical Marketing Reporter). 1986b. U.S. Imports of Chemicals and
Related Materials. New York: Schnell Publishing, April 7.
Cohen EN, Hood W. 1969. Application of low-temperature autoradiography
studies of the uptake and metabolism of volatile anesthetics in the
mouse. Anesthesiology 30:306-314.
Coleman EC, Ho C, Chang SS.'1981. Isolation and identification of
volatile compounds from baked potatoes. J Agric Food Chem 29:42-48.
Coleman WE, Lingg RD, Melton RG, Kopfler FC. 1976. The occurrence of
volatile organics in five drinking water supplies using gas
chromatography/mass spectrometry. In: Keith L, ed. Analysis and
Identification of Organic Substances in Water. Ann Arbor MI: Ann Arbor
Science, pp. 305-327.
Condie LW, Smallwood CL, Laurie RD. 1983. Comparative renal and
hepatotoxicity of halomethanes: Bromodichloromethane, bromoform,
chloroform, dibromochloromethane, and methylene chloride. Drug Chem
Toxicol 6(6):564-578.
Cornish HH. 1975. Solvents and vapors. In: Cassarett L, Doull J, eds.
Toxicology, the Basic Science of Poisons. MacMillan Publishing, p. 503
(cited in EPA 1985a).
Culliford D, Hewitt HB. 1957. The influence of sex hormone status on the
susceptibility of mice to chloroform-induced necrosis of the renal
tubules. J Endocrinol 14:381-393.
Danielsson BR, Ghantous H, Dencker L. 1986. Distribution of chloroform
and methyl chloroform and their metabolites in pregnant mice. Biol Res
Pregnancy Perinatol 7(2):77-83.
-------
94 Section 10
Danni 0, Brossa 0, Burdino E, Mlllillo P, Ugazio G. 1981. Toxicity of
halogenated hydrocarbons In pretreated rats: An experimental model for
the study of integrated permissible limits of environmental poison. Inc
Arch Occup Environ Health 49:105-112 (cited in EPA 1985a).
Davidson J. 1988. Written communication to Murphy J, Health Effects ODW
EPA, from Davidson J, Chairman. Testing Priority Committee OTS
March 1. 1988.
Demi E, Oesterle D. 1985. Dose-dependent promoting activity of
chloroform in rat liver foci bioassay. Cancer Lett 29:29-63.
* Deringer MK, Dunn TB, Heston WE. 1953. Results of exposure of strain
C3H mice to chloroform. Proc Soc Exp Biol Med 83:474-479.
De Serres FJ, Ashby J, eds. 1981. Evaluation of short-term tests for
carcinogens. In: Progress in Mutation Research. Vol I. Elsevier/North
Holland (cited in EPA 1985a).
* DeSalva S, Volpe A, Leigh G, Regan T. 1975. Long-term safety studies
of a chloroform-containing dentifrice and mouth rinse in man Food
Cosmet Toxicol 13:529-532.
Deshon HD. 1979. Carbon tetrachloride. In: Grayson M, Eckroth D, eds
Kirk-Othmer Encyclopedia of Chemical Technology, 3rd ed. Vol 5 New
York: John Wiley and Sons, pp. 693-703.
Dewalle FB, Chian ESK. 1981. Detection of trace organics in well water
near a solid waste landfill. J Am Water Works Assoc 73:206-211.
Dilley JV, Chemoff N, Kay D, Winslow N, Newell GW. 1977. Inhalation
teratology studies of five chemicals in rats. Toxicol Appl Pharmacol
41:196.
Dilling WL, Tefertiller NB, Kallos GJ. 1975. Evaporation rates of
methylene chloride, chloroform, 1,1,1-trichloroethane,
trichloroethylene, tetrachloroethylene, and other chlorinated compounds
in dilute aqueous solutions. Environ Sci Technol 9(9):833-838.
Dilling W. 1977. Interphase transfer processes. II. Evaporation rates of
chloromethanes, ethanes, ethylenes, propanes, and propylenes from dilute
aqueous solution. Comparisons with theoretical predictions. Environ Sci
Technol 11:405-409.
Dimitriades B, Joshi SB. 1977. Application of reactivity criteria in
oxidant-related emission control in the USA. In: Dimitriades B, ed.
International Conference on Photochemical Oxidant Pollution and Its
Control. Research Triangle Park, NC: EPA. EPA-600/3-77-001B
pp. 705-711.
Docks EL, Krishna G. 1976. The role of glutathione in chloroform-induced
hepatotoxicity. Exp Mol Pathol 24:13-22.
-------
References 95
Dow Chemical Company. 1988. Comments of the Dow Chemical Company on
ATSDR's Toxicological Profile for Chloroform. Submitted by the Dow
Chemical Company, Midland, MI, to Georgi Jones, Director, Office of
External Affairs, ATSDR, Atlanta, GA.
Eisenreich SJ, Looney BB, Thornton JD. 1981. Airborne organic
contaminants of the Great Lakes ecosystem. Environ Sci Technol
15(l):30-38.
Ekstrom T, Hoegbers J, Jernstroem B. 1982. Induction of hepatomas in
mice by repeated oral administration of chloroform with observations or
sex differences. J Natl Cancer Inst 5:251-255 (cited in EPA 1985a).
Entz RC, Thomas KW, Diachenko GW. 1982. Residues of volatile halocarbons
in foods using headspace gas chromatography. J Agric Food Chem
30:846-849.
EPA. 1980. Ambient Water Quality Criteria for Chloroform. Office of
Water Regulations and Standards. Washington, DC: Environmental
Protection Agency. EPA 440/5-80-033. NTIS PB81-117442.
EPA. 1981. Treatment Techniques for Controlling Trihalomethanes in
Drinking Water. Cincinnati, OH: Municipal Environmental Research Lab.
EPA-600/2-81-156.
EPA. 1982a. Test Methods for Evaluating Solid Wastes: Physical/Chemical
Methods. 5W-846. 2nd ed. Office of Solid Wastes.
EPA. 1982b. Methods for Organic Chemical Analysis of Municipal and
Industrial Wastewater. Cincinnati, OH: Environmental Monitoring and
Support Laboratory. EPA-600/4-82-057.
EPA. 1985a. Health Assessment Document for Chloroform. Final report.
Washington, DC: Office of Health and Environmental Assessment. EPA-
600/8-84-004F. NTIS PB86-105004/XAB.
EPA. 1985b. Reference Values for Risk Assessment. First draft. ECAO-
CIN-477. Cincinnati, OH: Environmental Criteria and Assessment Office.
EPA. 1985c. Survey of Chloroform Emission Sources. Research Triangle
Park, NC: Office of Air Quality. EPA 450/3-85-026.
EPA. 1985d. Criteria Document for Radium in Drinking Water: Draft. April
1985. Washington, DC: Health Effects Branch, Criteria and Standards
Division (WH-550), Office of Drinking Water.
EPA. 1985e. Notification requirements; reportable quantities and
adjustments. Final rule and proposed rule. 40 CFR Parts 117 and 302. Fed
Regist 50(65):13481.
EPA. 1985f. Intent to list chloroform as a hazardous air pollutant. Fed
Regist 50(188):39626-39629.
-------
96 Section 10
EPA. L986a. Evaluation of the Potential CarcLnogenicity of Chloroform
Review draft. Prepared by Carcinogen Assessment Group, Washington, DC:
Office of Health and Environmental Assessment. OHEA-C-073-54, December
1986.
EPA 1986b. Guidelines for the health assessment of suspect
developmental toxicants. Fed Regist 51(185):34028-34040.
EPA. 1987a. EPA Contract Laboratory Program. Statement of Work for
Organics Analysis, Multi-Media, Multi-Concentration. Revised 8/87.
EPA. 1987c. Extremely Hazardous Substances List and Threshold Planning
Quantities; Emergency Planning and Release Notification Requirements.
Fed Regist 52(77):13378-13410.
EPA. 1987d. Toxic Chemical Release Reporting; Community Right-to-Know.
Fed Regist 52(107):21152-21179.
EPA. 1987e. Integrated Risk Information System (IRIS) reference dose
(RfD) for oral exposure for chloroform. On Line (Verification date
12/2/85). Cincinnati, OH: Office of Health and Environmental Assessment,
Environmental Criteria and Assessment Office
EPA. 1987b. Integrated Risk Information System (IRIS) risk estimate for
carcinogenicity for chloroform. On line. Input pending (Verification
date 8/26/87). Cincinnati, OH: Office of Health and Environmental
Assessment, Environmental Criteria and Assessment Office
EPA. 1988. Analysis of Clean Water Act Effluent Guidelines Pollutants.
Summary of the Chemicals Regulated by Industrial Point Source Category
40 CFR Parts 400-475. Draft. Prepared by Industrial Technology Division
(WH 552) Office of Water Regulations and Standards. Office of Water.
Washington, DC: EPA.
EPA-NIH (National Institute of Health). 1987. OHM-TADS (Oil and
Hazardous Materials Technical Assistance Data System). On line.
Eschenbrenner AB, Miller E. 1945a. Induction of hepatomas in mice by
repeated oral administration of chloroform with observations on sex
differences. J Natl Cancer Inst 5:251-255.
Eschenbrenner A3, Miller E. 1945b. Sex differences in kidney morphology
and chloroform necrosis. Science 102:302-303.
Feingold A, Holaday DA. 1977. The pharmacokinetics of metabolism of
inhalation anaesthetics. Br J Anaesth 49:155-162.
Ferrario JB, Lawler GC, DeLeon IR, Laseter JL. 1985. Volatile organic
pollutants in biota and sediments of Lake Pontchartrain. Bull Environ
Contam Toxicol 34(2):246-255.
-------
References 97
Flathman PE, Dahlgran JR. 1982. Correspondence on: Anaerobic degradation
of halogenated 1- and 2-carbon organic compounds. Environ Sci Techno1
16:130.
Foerster EH, Garriott JC. 1981. Analysis for volatile compounds in
biological samples. J Anal Toxicol 5:241-244.
Fry BJ, Taylor T, Hathway DE. 1972. Pulmonary elimination of chloroform
and its metabolites in man. Arch Int Pharmacodyn 196:98-111.
Fujii T. 1977. Direct aqueous injection gas chromatography-mass
spectrometry for analysis of organohalides in water at concentrations
below the parts per billion level. J Chromatogr 139:297-302.
Gettler AO. 1934. Hedicolegal aspects of deaths associated with
chloroform or ether. Am J Surg 26:168-174.
Gettler AO, Blume H. 1931. Chloroform in the brain, lungs, and liver.
Quantitative recovery and determination. Arch Pathol 11:554-560.
Gocke E, King MT, Eckhardt K, Wild D. 1981. Mutagenicity of cosmetics
ingredients licensed by the European communities. Mutat Res 90:91-109.
Goodman LS, Gilman A. 1980. The pharmacological basis of therapeutics.
6th ed. New York: MacMillan Publishing (cited in EPA 1985a).
Gopinath C, Ford EJH. 1975. The role of microsomal hydroxylases in the
modification of chloroform and carbon tetrachloride. Toxicol Appl
Pharmacol 63:281-291 (cited in EPA 1985a).
Gosselin RE, Hodge HC, Smith RP, Gleason MN. 1976. Clinical Toxicology
of Commercial Products. Acute Poisoning. 4th ed. Baltimore, HD:
Williams and Wilkins (cited in EPA 1980).
Gram TE, Okine LK, Gram RA. 1986. The metabolism of xenobiotics by
certain extrahepatic organs and its relation to toxicity. Ann Rev
Pharmacol Toxicol 26:259-291.
Hammerstrand K. 1976. Chloroform in drinking water. Varian Instrum Appl
10:2-4 (cited in IARC 1979).
Hansch C, Leo AJ. 1985. Medchem Project Issue 26. Claremont CA: Pomona
College.
Hawley GG. 1981. The Condensed Chemical Dictionary. 10th ed. New York:
Van Nostrand Reinhold, p. 237.
Harkov R, Kebbekus B, Bozzelli JU. Lioy PJ, Daisey J. 1984. Comparison
of selected volatile organic compounds during the summer and winter at
urban sites in New Jersey. Sci Total Environ 38:259-274.
Harris RA, Groh GI. 1985. Membrane disordering effects of anesthetics
are enhanced by gangliosides. Anesthesiology (USA) 62(2):115-119.
-------
98 Section 10
Helz GR, Hsu RY. 1978. Volatile chloro- and bromocarbons in coastal
waters. Limnol Oceanogr 23:858-869.
Heukelekian H, Rand MC. 1955. Biochemical oxygen demand of pure organic
compounds. J Water Pollut Control Assoc 29:1040-1053.
Hewitt HB. 1956. Renal necrosis in mice after accidental exposure to
chloroform. Br J Exp Pathol 37:32-39.
Hewitt WR, Miyajima H, Cote KG, Plaa GC. 1979. Acute alteration of
chloroform-induced hepato- and nephrotoxicity by Mirex and Kepone.
Toxicol Appl Phannacol 48:509-527 (cited in EPA 1985a).
Hewitt WR, Miyajima H, Cote, MG, Plaa GC. 1980. Acute alteration of
chloroform-induced hepato- and nephrotoxicity by n-hexane, methyl
n-butyl ketone and 2,5-hexanedione. Toxicol Appl Pharmacol 53:230-248
(cited in EPA 1985a).
Hewitt WR, Brown EM, Plaa GL. 1983. Relationship between the carbon
skeleton length of ketonic solvents and potentiation of chloroform-
induced hepatotoxicity in rats. Toxicol Lett 16:297-304 (cited in EPA
1985a).
* Heywood R, Sortwell RJ, Noel PRB, et al. 1979. Safety evaluation of
toothpaste containing chloroform. III. Long-term study in beagle does
J Environ Toxicol 2:835-851.
* Hill RN. 1978. Differential toxicity of chloroform in the mouse Ann
NY Acad Sci 298:170-175.
Hill RN, Clemson TL, Liu DK, Vesell ES, Johnson WD. 1975. Genetic
control of chloroform toxicity in mice. Science 190:159-160.
Hjelle JJ, Gordon AS, Peterson DR. 1982. Studies on carbon
tetrachlorideethanol interactions in mice. Toxicol Lett 10:17-24.
Hook JB, Smith JH. 1985. Biochemical mechanisms of nephrotoxicity.
Transplant Proc (USA) 17(4 Suppl 1):41-50.
Hutzler NJ, Crittenden JC, Oravitz JL, Schaepe PA. 1983. Groundwater
transport of chlorinated organic compounds. In: Proceedings of the 186th
National Meeting of the American Chemical Society 23:499-502.
IARC (International Agency for Research on Cancer). 1979. Chloroform.
In: Some Halogenated Hydrocarbons. IARC Monographs on the Evaluation of
the Carcinogenic Risk of Chemicals to Humans. Lyons, France: World
Health Organization, IARC Vol. 20, pp. 401-427.
IARC (International Agency for Research on Cancer). 1982. Chloroform.
IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals
to Humans. Lyon, France: World Health Organization, IARC Suppl 4.
pp. 87-88.
-------
References 99
IRPTC (International Register of Potentially Toxic Chemicals). 1987.
IRPTC Data Profile on Chloroform. Switzerland: United Nations
Environment Programme, April 1987.
Jakobson I, Vahlberg JE, Holmberg B, Johansson G. 1982. Uptake via the
blood and elimination of 10 organic solvents following epicutaneous
exposure to anesthetized guinea pigs. Toxicol Appl Pharmacol 63:181-187
* Jones WM. Margolis G. Stephen CR. 1958. Hepatotoxicity of Inhalation
anesthetic drugs. Anesthesiology 19:715.
* Jorgenson TA, Rushbrook CJ. 1980. Effects of Chloroform in the
Drinking Water of Rats and Mice: Ninety-Day Subacute Toxicicy Study
(Final report, Phase 1). EPA-600/1-80-030. NTIS PB80-219108.
* Jorgenson TA, Meierhenry EF, Rushbrook CJ. et al. 1985.
Carcinogenicity of chloroform in drinking water to male Osborne-Mendel
rats and female B6C3F1 mice. Fund Appl Toxicol (USA) 5(4):760-769.
Kasso WV, Veils MR. 1981. A survey of trihalomethanes in the drinking
water system of Murfreesboro, Tennessee, USA. Bull Environ Contain
Toxicol 27:295-302.
Kawamura K, Kaplan IR. 1983. Organic compounds in the rainwater of Los
Angeles. Environ Sci Techno1 17:497-501.
Kawasaki M. 1980. Experiences with the test scheme under the chemical
control law of Japan. An approach to structure-activity correlations
Ecotoxic Environ Safety 4:444-454.
Keith H, Garrison AW, Allen FR, et al. 1976. Identification of organic
compounds in drinking water from thirteen cities. In: Keith LH, ed.
Ident Anal Organic Pollut Water. Ann Arbor, MI: Ann Arbor Press,
pp. 329-373.
Kier LE, Brusick DL, Auletta AE, et al. 1986. The Salmonella
cyphimuri tun/mammalian microsomal assay. A report of the U.S.
Environmental Protection Agency Gene-Tox Program. Mutat Res 168:69-76,
180-184, 202, 206. 207, 216-218, 224-240.
v.
* Kimura ET, Ebert DM, Dodge BW. 1971. Acute toxicity and limits of
solvent residue for sixteen organic solvents. Toxicol Appl Pharmacol
19:699-704.
Kirkland DJ, Smith KL, Van Abbe NJ. 1981. Failure of chloroform to
induce chromosome damage or sister-chromatid exchanges in cultured human
lymphocytes and failure to induce reversion in EscherLchia coli. Food
Cosmet Toxicol 19:651-656.
Klaunig JE, Ruch RJ, Pereira MA. 1986. Carcinogenicity of chlorinated
methane and ethane compounds administered in drinking water to mice.
Environ Health Perspect 69:89-95.
-------
100 Section 10
Kluwe WM, Hook JB. 1978. Polybrominated blphenyl-induced potentiation of
chloroform toxictty. Toxicol Appl Pharmacol 45:861-869 (cited in EPA
1985a).
Kroneld R. 1986. Chloroform in cap water and human blood. Bull Environ
Concam Toxicol 36:477-483.
Krotoszynski B, Bruneau GM, O'Neill HJ. 1979. Measunt of chemical
inhalation exposure in urban populations in the presence of endogeneous
effluents. J Anal Toxicol 3:225-234.
Kurtz CM, Bennett JH. Shapiro HH. 1936. Electrocardiographic studies
during surgical anesthesia. J Am Med Assoc 106:434-441 (cited in EPA
1985a).
Kutob SD, Plaa GL. 1962. The effect of acute ethanol intoxication on
chloroform-induced liver damage. J Pharmacol Exp Teratol 135:245-251
(cited in EPA 1985a).
* Kylin B, Reichard H, Sumegi I, Yllner S. 1963. Hepatotoxicity of
inhaled trichloroethylene. tetrachloroethylene, and chloroform. Single
exposure. Acta Pharmacol Toxicol 20:16-26.
* Land PD, Owen EL, Linde HW. 1981. Morphologic changes in mouse
spermatozoa after exposure to inhalational anesthetics during early
spermatogenesis. Anesthesiology 54:53-56.
Lehmann KB, Flury FF. 1943. Chlorinated hydrocarbons. In: Lehman KB,
Flury FF, eds. Toxicology and Hygiene of Industrial Solvents. Baltimore,
MD: Williams and Wilkins, pp. 138-145 and 191-196 (cited in EPA 1985a)
* Lehmann KB, Hasewaga. 1910. Studies of the absorption of chlorinated
hydrocarbons in animals and humans. Arch Hyg 72:327 (FRG) (cited in EPA
1985a).
* Lehmann KB. Schmidt-Kehl L. 1936. The thirteen most important
chlorinated aliphatic hydrocarbons from the standpoint of industrial
hygiene. Arch Hyg 116:131-200 (FRG) (cited in NIOSH 1974, EPA 1985a).
Liang JC, Hsu TC, Henry JE. 1983. Cytogenetic assays for mitotic
poisons: The grasshopper embryo system for volatile liquids. Mutat Res
113:467-479.
Lovegren NV, Fisher GS, Legendre MG, Schuller WH. 1979. Volatile
constituents of dried legumes. J Agric Food Chem 27:851-853.
* Lundberg I, Ekdahl M, Kronevi T. Lidums V, Lundberg S. 1986. Relative
hepatotoxicity of some industrial solvents after intraperitoneal
injection or inhalation exposure to rats. Environ Res 40(2):411-420.
Mabey W, Mill T. 1978. Critical review of hydrolysis of organic
compounds in water under environmental conditions. J Phys Chem Ref Data
7:383-415.
-------
References 101
Masuda Y, Nakayama N. 1982. Protective effect of diethyldlthiocarbamate
and carbon disulflde against liver injury induced by various hepatotoxic
agents. Biochem Pharmacol 31:2713-2725 (cited in EPA 1985a).
Masuda Y, Nakayama N. 1983. Protective action of diethyldithiocarbamate
and carbon disulfide" against renal injury induced by chloroform in mice
Biochem Pharmacol 31:2713-2725 (cited in EPA 1985a).
McConnell G. Ferguson DM. Pearson CR. 1975. Chlorinated hydrocarbons and
the environment. Endeavor 34:13-18.
McMartin ND, O'Connor JA Jr, Kaminsky LS. 1981. Effect of differential
changes in rat hepatic and renal cytochrome P-450 concentrations on
hepatotoxicity and nephrotoxicity of chloroform. Res Commun Chem Pathol
Pharmacol 31:99-110.
Mink FL, Brown TJ, Rickabaugh J. 1986. Absorption, distribution, and
excretion of carbon-14 trihalomethanes in mice and rats.-Bull Environ
Contam Toxicol 37(5):752-758.
Mirsalis JC, Tyson CK, Butterworth BE. 1982. Detection of genotoxic
carcinogens in the in vivo--in vitro hepatocyte DNA repair assay.
Environ Mutagenesis 4:553-562.
* Moore DH, Chasseaud LF, Majeed SK, Prentice DE, Roe FJC, Van abbe NJ.
1982. The effect of dose and vehicle on early tissue damage and
regenerative activity after chloroform administration to mice. Food Chem
Toxicol 20:951-954.
Morimoto K, Koizumi A. 1983. Trihalomethanes-induced sister chromatid
exchanges in human lymphocytes in vitro and mouse bone marrow cells in
vivo. Environ Res 32(1):72-79.
Morris LEV. 1951. Chloroform in blood and respired atmosphere. In:
Chloroform, A Study After 100 Years. Madison, WI: University of
Wisconsin Press, pp. 95-119 (cited in EPA 1985a).
Munson AE, Sain LE, Sanders VM, et al. 1982. Toxicology of organic
drinking water contaminants: Trichloromethane, bromodichloromethane,
dibromochloromethane, and tribromoethane. Environ Health'Perspect
46:117-126.
* Murray FJ, Schwetz BA, McBride JG, Staples RE. 1979. Toxicity of
inhaled chloroform in pregnant mice and their offspring. Toxicol Appl
Pharmacol 50:515-522.
NAS (National Academy of Sciences). 1977. Drinking Water and Health.
Vol. 1. Washington, DC: National Academy of Sciences, pp. 713-718.
* NCI (National Cancer Institute). 1976. Report on Carcinogenesis
Bioassay of Chloroform. NTIS PB-264018.
-------
102 Section 10
Newberne PM, Welgert J, Kula N. 1979. Effects of dietary fat on hepat
mixed function oxidases and hepatocellular carcinoma induced by
aflatoxin Bl in rats. Cancer Res 39:3986-3991.
1C
Newberne PM, Decamargo JLV, Clarke A. 1982. Choline deficiency, partial
hepatectomy, and liver tumors in rats and mice. Toxicol Pathol 10:95-109
(cited in Jorgenson et al. 1985).
Nicholson AA, Heresz 0, Lemyk B. 1977. Determination of free and total
potential halofonns in drinking water. Anal Chem 49:814-819.
Nicholson BC, Maguire BP, Bursill DB. 1984. Henry's Law constants for
the trihalomethanes: Effects of water composition and temperature
Environ Sci Technol 18:518-521.
* NIOSH (National Institute for Occupational Safety and Health). 1974
Criteria for a Recommended Standard...Occupational Exposure to
Chloroform. NTIS PB-246-695 (cited in EPA 1985a).
NIOSH (National Institute for Occupational Safety and Health). 1987c
NIOSH Manual of Analytical Methods. 3rd ed. Vol. 2. Cincinnati, OH:
Department of Health and Human Services. DHHS (NIOSH) Publ 84-100
Revision 1, 8/15/87.
NIOSH (National Institute for Occupational Safety and Health). 1977.
National Occupational Hazard Survey. Vol. 14. Survey Analyses and
Supplemental Tables. Cincinnati, OH: Department of Health, Education
and Welfare, pp. 2864-2866 (cited in EPA 1985a).
NIOSH (National Institute for Occupational Safety and Health). 1984
Current Awareness File. Registry of Toxic Effects of Chemical Substances
(RTECS). Cincinnati, OH: National Institute for Occupational Safety and
Health.
NIOSH (National Institute for Occupational Safety and Health). 1987a.
RTECS (Registry of Toxic Effects of Chemical Substances). Chloroform On
line, January 1987.
NLM (National Library of Medicine). 1987. Hazardous Substance Data Bank
Chloroform Record No. 56 (computer printout).
NTIS (National Technical Information Service). 1987. Federal Research in
Progress. On line, March 1987 (Dialog File 265).
NTP (National Toxicology Program). 1986. Annual Plan for Fiscal Year
1986. Research Triangle Park, NC: National Toxicology Program. Public
Health Service, Department of Health and Human Services, pp. 49 50
181. NTP-86-086.
NYDEC (New York State Department of Environmental Conservation). 1985
Superfund and Contract Laboratory Protocol, January 1985. NYDEC.
pp. D-54 to D-95.
-------
References 102
Orth OS, Liebenow RR, Capps RT. 1951. III-The effect of chloroform on
the cardiovascular system. In: Waters RM, ed. Chloroform, A Study After
100 Years. Madison, WI: University of Wisconsin Press, pp. 39-75 (cicec
in EPA 1985a).
OSHA (Occupational Safety and Health Administration). 1985. Permissible
Exposure Limits. Code of Federal Regulations 29:1910.1000.
* Palmer AK, Street AE, Roe FJC, Worden AN. Van Abbe NJ. 1979. Safety
evaluation of toothpaste containing chloroform. II. Long-term studies ir
rats. J Environ Pathol Toxicol 2:821-833.
Parsons JS, Mitzner S. 1975. Gas chromatographic method for
concentration and analysis of traces of industrial organic pollutants ir.
environmental air and stacks. Environ Sci Techno1 9:1053-1058.
Paul BB, Rubinstein D. 1963. Metabolism of carbon tetrachloride and
chloroform by the rat. J Pharmacol Exp Ther 141:141-148.
Pellizzari ED, Hartwell TD, Harris BSH, et al. 1982. Purgeable organic
compounds in mother's milk. Bull Environ Contain Toxicol 28:322-328.
Peoples AJ, Pfaffenberger CD, Shafik TM, Enos HF. 1979. Determination of
volatile purgeable halogenated hydrocarbons in human adipose tissue and
blood serum. Bull Environ Contam Toxicol 23:244-249.
Pereira MA, Lin LC, Lippitt JM, Herren SL. 1982. Trihalomethanes as
initiators and promoters of carcinogenesis. Environ Health Perspect
46:151-156.
Pereira MA, Daniel FB, Lin E. 1985. Relationship between metabolism of
haloacetonitriles and chloroform and their carcinogenic activity. In:
Jolley RL, et al., ed. Water Chlorination. Vol. 5. Chemistry,
Environmental Impact and Health Effects. Chelsea, MI: Lewis Publishers.
Pfaffenberger CD, Peoples AJ, Enos HF. 1980. Distribution of volatile
halogenated organic compounds between rat blood serum and adipose
tissue. Intern J Environ Anal Chem 8:55-65.
',
* Phoon WH, Goh KT. Lee LT, Tan KT, Kwok SE. 1983. Toxic jaundice from
occupational exposure to chloroform. Med J Malaysia 30:31-34.
Pohl LR, Gillette JR. 1984. Determination of toxic pathways of
metabolism by deuterium substitution. Drug Metab Rev (USA)
15(7):1335-1351.
Pohl LR, George JW, Satoh H. 1984. Strain and sex differences in
chloroform-induced nephrotoxicity. Different rates of metabolism of
chloroform to phosgene by the mouse kidney. Drug Metab Dispos
12(3):304-308.
-------
104 Section 20
Premel-Cabic A. Cailleux A, Allain P. 1974. A gas chromatographic assay
of fifteen volatile organic solvents in blood. Clin Chim Acta 56:5-11
(French) (cited in IARC 1979).
Rao PSC, Hornsby AG, Jessup RE. 1985. Indices for ranking the potential
for pesticide contamination of groundwater. Soil Crop Sci Soc Fl Proc
44-1-8.
Rachbun RE, Tai DY. 1981. Technique for determining the volatilization
coefficients of priority pollutants in streams. Water Res 15:243-250.
Rehm RM, Anderson ME. Duletsky SA, Misenheimer DC, Rollius HF.' 1982.
Chloroform Materials Balance. Draft report. EPA Contract 68-02-3168.
Task 69 (cited in EPA 1985a).
Reitz RH, Quast JF, Scott WT, Watanabe PG, Behring OJ. 1980.
Pharmacokinetics and macromolecular effects of chloroform in rats and
mice. Implications for carcinogenic risk estimation. Water Chlorinat
Environ Impact Health Eff 3:983-993.
Reitz RH, Fox TR, Quast JF. 1982. Mechanistic considerations for
carcinogenic risk estimation: Chloroform. Environ Health Perspect
46:163-168.
Reunanen M, Kroneld R. 1982. Determination of volatile halocarbons in
raw and drinking water, human serum, and urine by electron capture GC.
J Chromatogr Sci 20:449-454.
Reynolds ES. 1967. Liver parenchymal cell injury. IV. Pattern of
incorporation of carbon and chlorine from carbon tetrachloride into
chemical constituents of liver in vivo. J Pharmacol Exp Ther
155:177-126.
Reynolds ES. 1977. Liver endoplasmic reticulum. Target site of
halomethane metabolism. Adv Exp Med Biol 84:117.
Reynolds ES, Yee AG. 1967. Liver parenchymal cell injury. V.
Relationships between patterns of chloromethane--l^C incorporation into
constituents of liver in vivo and cellular injury. Lab Invest
16:591-603.
* Reynolds ES, Trelnen RJ, Farrish HH, Moslen MT. 1984a. Relationships
between the pharmacokinetics of carbon tetrachloride conversion to
carbon dioxide and chloroform and liver injury. Arch Toxicol 7:303-306
(cited in EPA 1985a).
Reynolds ES, Treinen RJ, Farrish HH, Moslen MT. 1984b. Metabolism of
14C-carbon tetrachloride to exhaled, excreted, and bound metabolites.
Dose-response, time-course, and pharmacokinetics. Biochem Pharmacol
33:3363-3374.
-------
References 105
* Roe FJC, Palmer AAK, Worden AN, Van Abbe NJ. 1979. Safety evaluation
of toothpaste containing chloroform. I. Long-term studies in mice. J
Environ Toxicol 2:799-819.
Roberts PV. Schreiner J. Hopkins GD. 1982. Field study of organic water
quality changes during groundwater recharge in the Palo Alto Baylands.
Water Res 16:1025-1035.
Rosenthal SL. 1987. A review of the mutagenicity of chloroform. Environ
Molec Hutagenesis 10:211-226.
Rubinstein D, Kanics L. 1964. The conversion of carbon tetrachloride and
chloroform to carbon dioxide by rat liver homogenates. Can J Biochem
42:1577-1585.
Rudali G. 1967. Oncogenic activity of some halogenated hydrocarbons used
in therapeutics. UICC Monogr Ser 7:138-143 (cited in EPA 1985a).
Sato A, Nakajima T. 1984. Dietary carbohydrate-induced and ethanol-
induced alteration of the metabolism and toxicity of chemical
substances. Nutr Cancer 6:121-132 (cited in EPA 1985a).
Sato A, Nakajima T, Koyama Y. 1980. Effects of chronic ethanol
consumption on hepatic metabolism of aromatic and chlorinated
hydrocarbons in rats. Br J Indust Med 37:382-386 (cited in EPA 1985a).
Sato A, Nakajima T, Koyama Y. 1981. Dose-related effect of a single dose
of ethanol on the metabolism in rat liver of some aromatic and
chlorinated hydrocarbons. Toxicol Appl Pharmacol 60:8-15 (cited in EPA
1985a).
Sax NI. 1979. Dangerous properties of industrial materials. 5th ed. New
York: Van Nostrand Reinhold, p. 193 (cited in EPA 1985a).
Scholler KL. 1970. Modification of the effects of chloroform on the rat
liver. Br J Anaesth 42:603-605 (cited in EPA 1985a).
* Schroeder HG. 1965. Acute and delayed chloroform poisoning. A case
report. Br J Anaesth 37:972-975.
* Schwetz BA, Leong BKJ, Gehring PJ. 1974. Embryo and fetotoxicity of
inhaled chloroform in rats. Toxicol Appl Pharmacol 28:442-451.
Shubik P, Ritchie AL. 1953. Sensitivity of male DBA mice to the toxicity
of chloroform as a laboratory hazard. Science 17:285 (cited in EPA
1985a).
Simmon VF, Kauhanen K, Tardiff RG. 1977. Mutagenic activity of chemicals
identified in drinking water. In: Scott D, Bridges BA, Sobels FH, eds.
Progress in Genetic Toxicology. Elsevier/North Holland Press,
pp. 249-258.
-------
106 Section 10
Singh HB, Salas LJ, Stiles RE. 1982. Distribution of selected gaseous
organic mutagens and suspect carcinogens in ambient air. In: Proceedings
of the Annual Meeting of the Air Pollution Control Association
75(4):82-65.1.
Single/ JE. 1984. Water (municipal treatment). In: Grayson M. Eckroth D
eds. Kirk-Othmer Encyclopedia of Chemical Technology, 3rd ed. Vol 24
New York: John Wiley and Sons, pp. 385-406.
Sipes IG, Krishna G, Gillette JR. 1977. Bioactivation of carbon
tetrachloride, chloroform, and bromotrichloromethane. Role of cytochroire
P-450. Life Sci 20:1541-1548.
* Smith AA, Volpitto PO, Gramling ZW, DeVore MB, Classman AB. 1973.
Chloroform, halothane, and regional anesthesia. A comparative study
Anesth Analg (Cleveland) 52:1-11.
Smith JH, Hook JB. 1983. Mechanism of chloroform nephrotoxicity. II In
vitro evidence for renal metabolism of chloroform in mice. Toxicol ADD!
Pharmacol 70(3):480-485.
Smith JH, Hook JB. 1984. Mechanism of chloroform nephrotoxicity. III.
Renal and hepatic mlcrosomal metabolism of chloroform In mice Toxicol
Appl Pharmacol 73(3):511-524.
Smith JH, Malta K, Sleight SD, Hook JB. 1984. Effect of sex hormone
status on chloroform nephrotoxicity and renal mixed function oxidases in
mice. Toxicology 30(4):305-316.
Smyth HF, Carpenter CP. Well CS, Pozzani UC, Striegel JA. 1962. Range
finding toxiclty data: List VI. Am Ind Hyg Assoc J 23:95-107.
Sporstoel S, Urdal K, Drangsholt H, GJoes N. 1985. Description of a
method for automated determination of organic pollutants in water Inc J
Environ Anal Chem 21:129-138.
SRI (Stanford Research Institute) 1987. Directory of Chemical
Producers: United States of America. Menlo Park, CA: SRI International
Steward A, Allot PR, Cowles AL, Mapleson WW. 1973. Solubility
coefficients for Inhaled anaesthetics for water, oil, and biological
media. Br J Anaesth 45:282-293.
Stewart RD. Dodd HC, Erly DS, Holder BB. 1965. Diagnosis of solvent
poisoning. J Am Med Assoc 193:1097-1100.
Strachan WMJ, Edwards CJ. 1984. Organic pollutants In Lake Ontario. Adv
Environ Scl Technol 14:239-264.
Sturrock J. 1977. Lack of mutagenic effect of halothane or chloroform on
cultured cells using the azaguanine test system. Br J Anaesth
49:207-210.
-------
References 10"
Symons JM, Bellar TA, Carswell JK, et al. 1975. National Organic
Reconnaissance Survey for halogenated organics. J An Water Works Assoc
67:634-647.
Tabak HH, Quave SA, Hashni CI, Barth EF. 1981. Biodegradability studies
with organic priority pollutant compounds. J Water Pollut Control Fed
53:1503-1518.
Taylor DC, Brown DM, Kuble R, Langley PF. 1974. Metabolism of
chloroform. II. A sex difference in the metabolism of ^C-chloroform ir
mice. Xenobiotica 4:165-174.
* Thompson DJ, Warnet SD, Robinson W. 1974. Teratology studies on
orally administered chloroform in the rat and rabbit. Toxicol Appl
Pharmacol 29:348-357.
Topham JC. 1980. Do induced sperm-head abnormalities in mice
specifically identify mammalian mutagens rather than carcinogens? Mucac
Res 74:379-387.
* Torkelson TR, Oyen F, Rove VK. 1976. The toxicity of chloroform as
determined by single and repeated exposure of laboratory animals. Am Ind
Hyg Assoc J 37:697-704.
Tsurata H. 1975. Percutaneous absorption of organic solvents. 1.
Comparative study of the in vivo percutaneous absorption of chlorinated
solvents in mice. Ind Health 13:227-236.
Tsurata H. 1977. Percutaneous absorption of organic solvents. 2. A
method for measuring the penetration rate of chlorinated solvents
through excised rat skin. Ind Health 15:131-140.
Tumasonis CF, McMartin DN, Bush B. 1985. Lifetime toxicity of chloroform
and bromodichloromethane when administered over a lifetime in rats.
Ecotoxicol Environ Saf 9(2):233-240.
Uehleke H,.Werner T. 1975. A comparative study on the irreversible
binding of labeled halothane, trichlorofluoromethane, chloroform, and
carbon tetrachloride to hepatic protein and lipids in vitro and in vivo
Arch Toxicol 34:289-303 (cited in EPA 1985).
Uehleke H, Werner T, Greim H, Kramer M. 1977. Metabolic activation of
halothane and tests in vitro for mutagenicity. Xenobiotica 7:393-400
USITC (United States International Trade Commission). 1986. Synthetic
Organic Chemicals, United States Production and Sales, 1985. USITC
Publication 1892. Washington DC: USITC.
Van Dyke RA, Chenoweth MB, Van Poznak A. 1964. Metabolism of volatile
anaesthetics. I. Conversion in vivo of several anesthetics to ^C02 and
chloride. Blochem Pharmacol 13:1239-1247.
-------
108 Section 10
Varma MM, Balram A. Katz HM. 1984. Trihalomethanes in groundwater
systems. J Environ Syst 14(2):115-126.
Veiro JA, Hunt GRA. 1985. The modulation of ion channels by the
inhalation of general anaesthetics. A supplemental 1H-NMR investigation
using unilamellar phospholipid membranes Chem Biol Interact (Ireland)
54(3)-337-348.
Von Oettingen WF. 1955. The halogenated hydrocarbons toxicity and
potential dangers. Washington, DC: U.S. Department of Health, Education,
and Welfare, Government Printing Office (cited in EPA 1980).
Von Oettingen WF. 1964. The Halogenated Hydrocarbons of Industrial and
Toxicological Importance. Amsterdam: Elsevier, pp. 77-108 (cited in EPA
1985a).
* Wallace CJ. 1950. Hepatitis and nephrosis due to cough syrup
containing chloroform. Calif Med 73:442.
Wallace LA. 1986. Personal exposures, indoor and outdoor air
concentrations, and exhaled breath concentrations of selected volatile
organic compounds measured for 600 residents of New Jersey, North
Dakota, North Carolina, and California. Toxicol Environ Chem
612:215-236.
Wallace L, Pellizzari E, Hartwell T, et al. 1986a. Concentrations of 20
volatile organic compounds in the air and drinking water of 350
residents of New Jersey compared with concentrations in their exhaled
breath. J Occup Med 28:603-607.
Wallace L, Pellizzari E, Sheldon L, Hartwell T, Sparacino C, Zelon H
1986b. The total exposure assessment methodology (TEAM) study: Direct
measurement of personal exposures through air and water for 600
residents of several cities. In: Cohen Y, ed. Pollutants in Multimedia
Environment. Plenum Publishing, pp. 289-315.
Wallace L, Pellizzari E, Leaderer B, Zelon H, Sheldon L. 1987. Emissions
of volatile organic compounds from building material and consumer
products. Atmos Environ 21:385-395.
White AE, Takehisa S, Eger El, Wolff S, Stevens WC. 1979. Sister
chromatid exchanges induced by inhaled anesthetics. Anesthesiology
50:426-430 (cited in EPA 1985a).
Wilson J, Enfield CG, Dunlap VJ, Cosby RL, Foster DA, Baskin LB. 1981.
Transport and fate of selected organic pollutants in a sandy soil. J
Environ Qual 10:501-506.
Wilson JT, MeNabb JF, Wilson BH, Noonan MJ. 1983. Biotransformation of
selected organic pollutants in groundwater. Dev Ind Microbiol
24:225-233.
-------
References 109
Windholz M, ed. 1983. The Merck Index. 10th ed. Rahway, NJ: Merck and
Co., pp. 300-301.
Withey JR. Collins BT, Collins PG. 1983. Effect of vehicle on the
pharmacokinetics and uptake of four halogenated hydrocarbons from the
gastrointestinal tract of the rat. J Appl Toxicol 3:249-253.
Wolf CR. Mansuy D, Nastainczyk W, Deutschmann G. Ullrich V. 1977. The
reduction of polyhalogenated methanes by liver microsomal cytochrome
P-450. Mol Pharmacol 13:698-705 (cited in EPA 1985a).
Wood-Smith FG, Stewart HC. 1964. Drugs in Anesthetic Practice.
Washington, DC: Butterworth, pp. 131-135 (cited in EPA 1985a).
Young TB, Kanarek MS, Tslatis AA. 1981. Epidemiologic study of drinking
water chlorination, Wisconsin female cancer mortality. J Natl Cancer
Inst 67(6):1191-1198.
Zeller A. 1883. On the fate of iodoforms and chloroform in the organism
Hoppe-Seyler's Z Physiol Chem 8:78-79 (cited in EPA 1985a).
Zoeteman BCJ, Harmsen K, Linders JBHJ, Morra CFH, Sloof W. 1980.
Persistent organic pollutants in river water and groundwater of the
Netherlands. Chemosphere 9:231-249.
-------
ILL
11. GLOSSARY
Acute Exposure--Exposure to a chemical for a duration of 14 days or
less, as specified in the Toxicological Profiles.
Bioconcentration Factor (BCF)--The quotient of the concentration of a
chemical in aquatic organisms at a specific time or during a discrete
time period of exposure divided by the concentration in the surrounding
water at the same time or during the same time period.
Carcinogen--A chemical capable of inducing cancer.
Ceiling value (CL)--A concentration of a substance that should not be
exceeded, even instantaneously.
Chronic Exposure--Exposure to a chemical for 365 days or more, as
specified in the Toxicological Profiles.
Developmental Toxicity--The occurrence of adverse effects on the
developing organism that may result from exposure to a chemical prior co
conception (either parent), during prenatal development, or postnatally
to the time of sexual maturation. Adverse developmental effects may be
detected at any point in the life span of the organism.
Embryotoxicity and Fetotoxicity--Any toxic effect on the conceptus as a
result of prenatal exposure to a chemical; the distinguishing feature
between the two terms is the stage of development during which the
insult occurred. The terms, as used here, include malformations and
variations, altered growth, and in utero death.
Frank Effect Level (FEL)--That level of exposure which produces a
statistically or biologically significant increase in frequency or
severity of unmistakable adverse effects, such as irreversible
functional impairment or mortality, in an exposed population when
compared with its appropriate control.
EPA Health Advisory--An estimate of acceptable drinking water levels for
a chemical substance based on health effects information. A health
advisory is not a legally enforceable federal standard, but serves as
technical guidance to assist federal, state, and local officials.
Immediately Dangerous to Life or Health (IDLH)--The maximum
environmental concentration of a contaminant from which one could escape
within 30 min without any escape-impairing symptoms or irreversible
health effects.
-------
112 Section 11
Intermediate Exposure--Exposure to a chemical for a duration of 15-36u
days, as specified in the Toxicological Profiles.
Immunologic Toxicity--The occurrence of adverse effects on the immune
system that may result from exposure to environmental agents such as
chemicals.
In vitro--Isolated from the living organism and artificially maincair.ec.
as in a test tube.
In vivo--Occurring within the living organism.
Key Study--An animal or human toxicological study that best illustrates
the nature of the adverse effects produced and the doses associated with
those effects.
Lethal Concentration(LO) (LCLO)--The lowest concentration of a chemical
in air which has been reported to have caused death in humans or
animals.
Lethal Concentration^)
-------
Glossary 113
Neurotoxicity--The occurrence of adverse effects on the nervous system
following exposure to a chemical.
No -Observed -Adverse -Effect Level (NOAEL)--That dose of chemical at which
there are no statistically or biologically significant increases in
frequency or severity of adverse effects seen between the exposed
population and its appropriate control. Effects may be produced at this
dose, but they are not considered to be adverse.
No-Observed-Effect Level (NOEL) --That dose of chemical at which there
are no statistically or biologically significant increases in frequency
or severity of effects seen between the exposed population and its
appropriate control.
Permissible Exposure Limit (PEL) --An allowable exposure level in
workplace air averaged over an 8-h shift.
q^ --The upper-bound estimate of the low-dose slope of the dose-response
curve as determined by the multistage procedure. The q * can be used to
calculate an estimate of carcinogenic potency, the incremental excess
cancer risk per unit of exposure (usually /ig/L for water, mg/kg/day for
food, and Mg/n^ for air) .
Reference Dose (RfD)--An estimate (with uncertainty spanning perhaps an
order of magnitude) of the daily exposure of the human population to a
potential hazard that is likely to be without risk of deleterious
effects during a lifetime. The RfD is operationally derived from the
NOAEL (from animal and human studies) by a consistent application of
uncertainty factors that reflect various types of data used to estimate
RfDs and an additional modifying factor, which is based on a
professional judgment of the entire database on the chemical. The RfDs
are not applicable to nonthreshold effects such as cancer.
Reportable Quantity (RQ)--The quantity of a hazardous substance that is
considered reportable under CERCLA. Reportable quantities are: (1) 1 Ib
or greater or (2) for selected substances, an amount established by
regulation either under CERCLA or under Sect. 311 of the Clean Water
Act. Quantities are measured over a 24 -h period.
Reproductive Toxiclty--The occurrence of adverse effects on the
reproductive system that may result from exposure to a chemical. The
toxic ity may be directed to the reproductive organs and/or the related
endocrine system. The manifestation of such toxicity may be noted as
alterations in sexual behavior, fertility, pregnancy outcomes, or
modifications in other functions that are dependent on the integrity of
this system.
Short-Term Exposure Limit (STEL)--The maximum concentration to which
workers can be exposed for up to 15 min continually. No more than four
excursions are allowed per day, and there must be at least 60 min
between exposure periods. The daily TLV-TWA may not be exceeded.
-------
114 Section 11
Target Organ Toxlcity--This term covers a broad range of adverse effects
on target organs or physiological systems (e.g., renal, cardiovascular)
extending from those arising through a single limited exposure to those
assumed over a lifetime pf exposure to a chemical.
Teratogen--A chemical that causes structural defects that affect the
development of an organism.
Threshold Limit Value (TLV)--A concentration of a substance to which
most workers can be exposed without adverse effect. The TLV may be
expressed as a TWA, as a STEL, or as a CL.
Time-weighted Average (TWA)--An allowable exposure concentration
averaged over a normal 8-h workday or 40-h workweek.
Uncertainty Factor (UF)--A factor used in operationally deriving the RfD
from experimental data. UFs are intended to account for (1) the
variation in sensitivity among the members of the human population,
(2) the uncertainty in extrapolating animal data to the case of humans.
(3) the uncertainty in extrapolating from data obtained in a study that
is of less than lifetime exposure, and (4) the uncertainty in using
LOAEL data rather than NOAEL data. Usually each of these factors is set
equal to 10.
-------
LL5
APPENDIX: PEER REVIEW
A peer review panel was assembled for chloroform. The panel
consisted of the following 'members: Dr. Herbert Cornish, (retired)
Professor of Toxicology, University of Michigan; Dr. Derek Hodgson,
Vice Chairman, Department of Chemistry, University of North Carolina,
Chapel Hill; and Dr. Richard Bull, Associate Professor of
Pharmacology/Toxicology, College of Pharmacy, Washington State
University. These experts collectively have knowledge of chloroform's
physical and chemical properties, toxicokinetics, key health end points,
mechanisms of action, human and animal exposure, and quantification of
risk to humans. All reviewers were selected in conformity with the
conditions for peer review specified in the Superfund Amendments and
Reauthorization Act of 1986, Section 110.
A joint panel of scientists from ATSDR and EPA has reviewed the
peer reviewers' comments and determined which comments will be included
in the profile. A listing of the peer reviewers' comments not
incorporated in the profile, with a brief explanation of the rationale
for their exclusion, exists as part of the administrative record for
this compound. A list of databases reviewed and a list of unpublished
documents cited are also included in the administrative record.
The citation of the peer review panel should not be understood to
imply their approval of the profile's final content. The responsibility
for the content of this profile lies with the Agency for Toxic
Substances and Disease Registry.
------- |