EPA
United States
Environmental Protection
Agency
HEALTH RISKS TO
FETUSES, INFANTS
AND CHILDREN
(PROPOSED STAGE 2
DISINFECTANT/
DISINFECTION
BYPRODUCTS): A
REVIEW
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OST (Mail Code 4304-T)
EP A-822-R-03 -010
http://www.epa.gov/waterscience/humanhealth/
March, 2003
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Table of Contents
Table of Contents iii
LIST OF TABLES vi
ABBREVIATIONS \ iii
EXECUTIVE SUMMARY ix
PLAIN LANGUAGE SUMMARY OF EVALUATION OF CHILDREN'S RISK xxxiv
1. INTRODUCTION 1
1.1. RISK TO CHILDREN 1
1.2. RISK ASSESSMENT METHODS 6
1.3 DETERMINING RISK TO CHILDREN 14
1.4. MAXIMUM CONTAMINANT LEVEL GOAL AND MAXIMUM RESIDUAL
DISINFECTANT LEVEL GOAL 17
2. ESTIMATES OF RISK TO CHILDREN FOR STAGE II DISINFECTANTS/
DISINFECTION BYPRODUCTS 25
2.1. CHLORINATED DRINKING WATER 25
2.1.1. Developmental/Reproductive Effects 26
2.1.2. Systemic Effects 40
2.1.3. Carcinogenicity 40
2.2. TRIHALOMETHANES 42
2.2.1. Chloroform 42
2.2.1.1. Developmental/Reproductive Effects 42
2.2.1.2. Systemic Effects 48
2.2.1.3. Carcinogenicity 48
2.2.1.4. Basis for RfD and MCLG 53
2.2.1.5. Children's Risk in Relation to the MCLG 55
2.2.2. Brominated Trihalomethanes 56
2.2.2.1. Bromodichloromethane 64
Developmental/Reproductive Effects 64
Systemic Effects 74
Carcinogenicity 74
Basis for RfD and MCLG 77
Children's Risk in Relation to the MCLG 77
2.2.2.2. Dibromochloromethane 78
Developmental/Reproductive Effects 78
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Systemic Effects 79
Carcinogenicity 80
Basis for RfD and MCLG 83
Children's Risk in Relation to the MCLG 83
2.2.2.3. Bromoform 84
Developmental/Reproductive Effects 84
Systemic Effects 86
Carcinogenicity 87
Basis for RfD and MCLG 89
Children's Risk in Relation to the MCLG 90
2.3. HALOACETIC ACIDS 90
2.3.1. Monochloroacetic Acid 92
Developmental/Reproductive Effects 92
Systemic Effects 95
Carcinogenicity 95
Basis for RfD and MCLG 97
Children's Risk in Relation to the MCLG 97
2.3.2. Dichloroacetic Acid 98
Developmental/Reproductive Effects 98
Systemic Effects 102
Carcinogenicity 106
Basis for RfD and MCLG 108
Children's Risk in Relation to the MCLG 109
2.3.3. Trichloroacetic Acid Ill
Developmental/Reproductive Effects Ill
Systemic Effects 113
Carcinogenicity 114
Basis for the RfD and MCLG 117
Children's Risk in Relation to the MCLG 118
2.3.4. Monobromoacetic Acid 118
Developmental/Reproductive Effects 118
Systemic Effects 120
Carcinogenicity 120
Basis for RfD and MCLG 121
Children's Risk in Relation to the MCLG 121
2.3.5. Bromochloroacetic Acid 121
Developmental/Reproductive Effects 121
Systemic Effects 126
Carcinogenicity 127
Basis for RfD and MCLG 127
Children's Risk in Relation to the MCLG 127
2.3.6. Dibromoacetic Acid 128
Developmental/Reproductive Effects 128
Systemic Effects 138
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Carcinogenicity 140
Basis for RfD and MCLG 141
Children's Risk in Relation to the MCLG 142
2.4. BROMATE 143
2.4.1. Developmental/Reproductive Effects 143
2.4.2. Systemic Toxicity 145
2.4.3. Carcinogenicity 148
2.4.4. Basis for the RfD and MCLG 150
2.4.5. Children's Risk in Relation to the MCLG 150
2.5. CHLORITE/CHLORINE DIOXIDE 151
2.5.1. Developmental/Reproductive Effects 152
2.5.2. Systemic Toxicity 159
2.5.3. Carcinogenicity 161
2.5.4. Basis for RfD, MCLG, and MRDLG 163
2.5.5. Children's Risk in Relation to the MCLG and MRDLG 163
2.6. CHLORINE 164
2.6.1. Developmental/Reproductive Effects 165
2.6.2. Systemic Toxicity 167
2.6.3. Carcinogenicity 167
2.6.4. Basis for RfD and MRDLG 168
2.6.5. Children's Risk Relative to the MRDLG 168
2.7. CHLORAMINE 169
2.7.1. Developmental/Reproductive Effects 169
2.7.2. Systemic Toxicity 169
2.7.3. Carcinogenicity 171
2.7.4. Basis for the RfD and MRDLG 171
2.7.5. Children's Risk in Relation to the MRDLG 172
2.8. MX [3-Chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone] 173
2.8.1. Developmental/Reproductive Effects 173
2.8.2. Systemic Effects 175
2.8.3. Carcinogenicity 175
2.8.4. Basis for RfD and MCLG 177
2.8.5. Children's Risk in Relation to the MCLG Ill
3. SUMMARY AND CONCLUSIONS 178
4. REFERENCES 181
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LIST OF TABLES
Table ES-1. Comparison of Toxicity Endpoints xxix
Table 1. Disinfectant Byproducts and their MCLGs Considered in this Document 13
Table 2. Disinfectants and their MRDLGs Considered in this Document 13
Table 3. Comparison of Toxicity Endpoints 14
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This document is an update of Health Risks to Fetuses, Infants and Children (Final Stage 1
D/DBP Rule, 1998). The new information that has been added to this document is based in large
part on the contents of the Drinking Water Criteria Documents (CDs) for Stage 2 disinfection
byproducts that were available in the Office of Science and Technology as of March, 2003:
Contributors
Nancy Chiu, Ph.D.
U.S.
EPA
Office
of Science
and
Technology,
Office
of Water
Diana Wong, Ph.D., DABT
U.S.
EPA
Office
of Science
and
Technology,
Office
of Water
Julie Du, Ph.D.
U.S.
EPA
Office
of Science
and
Technology,
Office
of Water
Mary Ko Manibusan, M.P.H.
U.S.
EPA
Office
of Science
and
Technology,
Office
of Water
Ambika Bathija, Ph.D.U.S. EPA Office of Science and Technology, Office of Water
Joyce Donohue, Ph.D. U.S. EPA Office of Science and Technology, Office of Water
Amal M. Mahfouz, Ph.D. U.S. EPA Office of Science and Technology, Office of Water
Rita Schoeny, Ph.D. U.S. EPA Office of Science and Technology, Office of Water
*Hend Galal-Gorchev, Ph.D. World Health Organization (WHO) Coordinator (retired), Geneva,
Switzerland
Peer Reviews
This document is based on Drinking Water Criteria Documents that have been peer reviewed.
The epidemiology information used in this document was also peer reviewed by Pauline Mendola,
Ph.D., U.S. EPA Office of Research and Development.
* Dr. Galal-Gorchev helped the Health and Ecological Criteria Division's staff (HECD) in the
development of the Criteria Document on dichloroacetic acid (DC A).
This document was prepared under the U.S. EPA contract No. 68-C-99-232 with GRAM, Inc.,
and Toxicology Excellence for Risk Assessment (TERA) under the lead of Lynne Haber, and U.S.
EPA contract No. 68-C-09-026 with the CADMUS Group and TERA, under the lead of Qiyu
(Jay) Zhao, respectively.
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ABBREVIATIONS
BBDR
biologically based dose response
BCA
bromochloroacetic acid
BDCM
bromodichloromethane
BMD
benchmark dose
BMDL
lower confidence limit on benchmark dose
CI
confidence interval
CYP
cytochrome P450
DBA
dibromoacetic acid
DBCM
dibromochloromethane
DBP
disinfection byproduct
DCA
dichloroacetic acid
D/DBP
disinfectant and disinfection byproduct
DWEL
drinking water equivalent level
GD
gestation day
GLP
good laboratory practice
GST
glutathione S-transferase
hCG
human chorionic gonadotropin
IRIS
Integrated Risk Information System
LH
luteinizing hormone
LHRH
luteinizing hormone releasing hormone
LOAEL
lowest-observed-adverse-effect level
MBA
monobromoacetic acid
MCA
monochloroacetic acid
MCLG
maximum contaminant level goal
MCL
maximum contaminant level
MF
modifying factor
MRDLG
maximum residual disinfectant level goal
MX
3-Chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone
NOAEL
no-observed-adverse-effect level
OR
odds ratio
pnd
postnatal day
POR
prevalence odds ratio
R£D
reference dose
RSC
relative source contribution
TCA
trichloroacetic acid
THM
trihalomethane
TTHM
total THM
UF
uncertainty factor
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EXECUTIVE SUMMARY
EPA's Office of Water is charged with ensuring that the United States population has
clean water and safe drinking water. Untreated water can be contaminated with microbial agents
that can cause infectious diseases. The most common result of infection by waterborne pathogens
is diarrheal disease, which may have serious consequences for children, the elderly, pregnant
women, and people of all ages with immune deficiencies. Water is often treated with chemical
disinfectants to prevent waterborne infectious disease; this process results in formation of
chemical disinfection byproducts.
EPA's drinking water assessments and regulations have historically considered sensitive
subpopulations that may be at increased risk. EPA's policy, the Safe Drinking Water Act
Amendments of 1996, and Executive Order 13045 call for the consideration of sensitive
populations, with particular attention to fetuses, infants and children in conducting risk
assessments and characterizations, and in establishing regulations and public health standards.
This document considers the potential risks to children1 from 13 disinfection byproducts (DBPs)
and three disinfectants. Establishment of a Maximum Contaminant Level Goal (MCLG) for a
specific contaminant is based on available evidence of carcinogenicity or noncancer adverse health
effects from drinking water exposure using the EPA's guidelines for risk assessment (see 63 FR
69390 for a detailed discussion of the process for establishing MCLGs). MCLGs are
nonenforceable health goals. As with MCLGs, Maximum Residual Disinfectant Level Goal
(MRDLGs) are nonenforceable health goals for drinking water disinfectants, based only on health
'For the purposes of this document, effects on children are defined as effects on a parent's
reproductive capacity, and effects on the fetus, infant, and children through sexual and physical
maturity (typically at age 18).
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effects and exposure information, and are established at the level at which no known or
anticipated adverse effects on the health of persons occur, and which allows an adequate margin
of safety. The MCLGs and MRDLGs are used as the basis for setting the Maximum Contaminant
Levels (MCLs), which are the enforceable drinking water standards. MCLs are set as close to the
MCLGs and MRDLGs as feasible, taking costs, treatment technology, and other considerations
into account.
The proposed Stage 2 Disinfectant/Disinfection Byproduct (D/DBP) Rule proposes new
MCLGs for three DBPs: monochloroacetic acid (MCA), trichloroacetic acid (TCA), and
chloroform. MCLGs and MRDLGs promulgated in 1998 for six DBPs and three disinfectants are
also discussed in this document. Insufficient data were available to develop MCLGs for four
DBPs.
This document updates the document on health risks to fetuses, infants, and children that
was prepared for the final Stage 1 D/DBP Rule (EPA, 1998c), and includes a number of new
developmental and reproductive toxicity studies in animals, new epidemiology studies of
developmental/reproductive effects, new cancer studies in animals, and the derivation of several
new risk values. To evaluate the toxicity of disinfectants and disinfection byproducts (D/DBPs)
for fetuses, infants, and children, EPA asked the following questions in deciding that the MCLG
or MRDLG for each D/DBP is protective of children:
1. Is there information that shows that the D/DBP causes effects in the developing
fetus or harms a woman's ability to become pregnant and bear children? If it
causes these effects, do these effects occur at lower doses than those that cause
other types of effects?
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2. If the D/DBP causes cancer, are children more likely to be affected by a given dose
than are adults?
3. If the D/DBP causes a health effect other than cancer, such as an effect on the liver
or kidney, are children affected at lower doses than are adults?
All available data were considered in addressing these questions, including any data
directly comparing toxicity of the D/DBP in children and adults, and any available mechanistic
data. Based on an evaluation of the data for each D/DBP, it can be concluded that the
MCLGs/MRDLGs of all the D/DBPs in the proposed Stage 2 D/DBP Rule are protective of
fetuses, infants and children. This conclusion is based on the data summarized in this document;
data on each DBP are described in detail in the supporting criteria documents. There are
uncertainties in this conclusion, however, due to incomplete information on the sensitivity of
children to systemic effects, and on age-related changes in the metabolism of the D/DBPs.
Chlorinated Drinking Water. There are no reliable animal studies on the reproductive
or developmental toxicity of chlorinated drinking water, nor are there animal studies addressing
age-related differences in the systemic response to chlorinated drinking water, or in the potential
carcinogenicity of chlorinated drinking water.
Epidemiology studies suggest that DBPs are associated with developmental and
reproductive effects under certain exposure conditions. The existing data are still relatively sparse,
and are insufficient for dose-response analysis. There are inconsistencies among the available
studies on the association between drinking water disinfection and specific effects, such as
changes in the menstrual cycle, fetal growth, fetal viability, and congenital abnormalities. The
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studies with the best exposure assessment found the strongest association between changes in
fetal growth and fetal viability and chlorinated drinking water exposure. Uncertainty in evaluating
the effects of DBPs is enhanced by the difficulties in exposure assessment, including the different
composition of D/DBPs at different locations within a water system, and the variation in
composition over time at a single location. The specific DBP(s) responsible for reproductive
and/or developmental effects is not known. The existing data suggest a relationship between
THMs, particularly BDCM, and developmental effects. This may be related to a tendency among
investigators to report concentrations of THMs, but not the concentrations of other DBPs. The
available data suggest that additional studies on THMs and other DBPs are needed.
Chloroform. A prospective epidemiology study found no clear association between
chloroform exposure and reduced menstrual cycle length. Studies in rats, rabbits and mice have
reported developmental effects, including pup weight reduction, teratogenicity and
embryolethality, from chloroform administration (Murray et al., 1979; Ruddick et al., 1983;
Schwetz et al., 1974). These prenatal effects, however, were typically associated with exposures
causing maternal toxicity, and occurred at oral doses above those causing hepatotoxicity. In a
multigeneration reproductive assay in CD-I mice treated with chloroform (NTP, 1988), no
adverse effects on fertility or reproduction were observed, although increased liver weight and
liver lesions were observed in treated females.
Animal studies in rats, mice, and dogs have shown that the liver and kidney are the main
target organs for chloroform. The current RfD for chloroform was estimated to be 0.01
mg/kg/day based on hepatotoxicity (fatty cysts and an increase of serum glutamic pyruvic
transaminase) observed in an oral study in dogs (Heywood et al., 1979).
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Chloroform has also been found to cause liver and kidney tumors in rodents (EPA, 1994a,
1998a). Under EPA's Draft Guidelines for Carcinogen Risk Assessment (EPA, 1999),
chloroform is likely to be carcinogenic to humans by all routes of exposure under high exposure
conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues (EPA,
2001a). Chloroform is not likely to be carcinogenic to humans by any route of exposure under
exposure conditions that do not cause cytotoxicity and cell regeneration. This weight-of-
evidence conclusion is based on observations in animals exposed to chloroform. The carcinogenic
potential of chloroform was evaluated using a nonlinear approach, in which liver toxicity was
considered the most sensitive effect for chloroform and as a key event in its carcinogenicity.
The MCLG is based on an oral study in adult dogs (Heywood et al., 1979), which was
used to derive the RfD (EPA, 2001a). This RfD corresponds to an MCLG of 0.07 mg/L, based
on a 70 kg adult consuming 2 liters of water per day (a general assumption used in this document)
and a 20% relative source contribution (RSC) from drinking water (EPA, 200Id). The MCLG is
considered protective of both adults and children based on several lines of reasoning. First,
developmental effects occurred at doses above those causing hepatotoxicity. Second, the weight
of evidence indicates that the mode of action by which chloroform produces organ toxicity and
carcinogenicity is the same for children and adults. Therefore, for both children and adults,
protection from organ toxicity would also provide protection from the carcinogenic effects of
chloroform. Finally, the available data indicate that children are not uniquely sensitive to the
organ toxicity caused by high doses of chloroform and there is no evidence from the available
studies to suggest that children or fetuses would be qualitatively more sensitive to its effects than
adults. For example, the liver toxicity observed in mice exposed prenatally, postnatally and as
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adults in a two-generation study was similar to the toxicity observed in mice exposed in a different
study for a similar duration, beginning shortly after weaning. This comparison is far from perfect,
but it allows a general comparison of the chloroform doses that cause liver effects in young mice
and adults. Based on these considerations, the MCLG based on systemic effects in adult dogs
protects children from the reproductive, developmental, systemic, and carcinogenic effects of
chloroform.
Bromodichloromethane (BDCM). A prospective epidemiology study suggested that
increasing levels of BDCM were associated with significantly shorter menstrual cycles. Because
the subjects were also exposed to other contaminants and disinfection byproducts in the drinking
water, establishment of a causal relationship between BDCM exposure and the observed effect is
difficult.
Reproductive/developmental assays with BDCM in animals, and epidemiology studies
regarding these effects of BDCM have shown mixed results. Full litter resorptions have been
reported in some rat studies (Bielmeier et al., 2001; Narotsky et al., 1997a); the effect appears to
be specific to F344 rats and not Sprague-Dawley rats, and it is unclear whether the proposed
mechanism of action is relevant to humans. An increase in stillbirths or spontaneous abortion has
been reported in some epidemiological studies as being associated with increased BDCM levels
(King et al., 2000a; Waller et al., 1998), but the data are insufficient to show causality of BDCM.
Other studies in rats and rabbits reported mild developmental delay, possibly secondary to
decreased water consumption (CCC, 2000a, c; Ruddick et al., 1983; Christian et al., 2002), or did
not report developmental or reproductive effects (CCC, 2000b; NTP, 1998a); all tested up to
maternally toxic doses.
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The liver and kidney are the main targets in rats of systemic effects of acute and chronic
BDCM exposure. Effects on the liver, kidney and thyroid have been noted in male mice (Aida et
al., 1992; NTP, 1987). EPA has calculated an RfD of 0.002 mg/kg/day based on a LOAEL of
6.1 mg/kg/day in a chronic study (Aida et al., 1992), in which the critical endpoint was liver fatty
degeneration and granuloma in male rats (EPA, 2002c). Due to the lack of a multigeneration
reproductive toxicity study and uncertainty related to possible reproductive or developmental
effects suggested by epidemiological studies (EPA, 2002c), a 3-fold uncertainty factor (UF) for
database deficiencies was used in the derivation of the RfD.
Studies have shown that BDCM exposure results in tumors of the large intestine and
kidneys in male and female rats, the kidneys in male mice and the liver in female mice (NTP,
1987). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, BDCM has been
classified as likely to be carcinogenic to humans, with a cancer oral slope factor of 8.1 x 10"3 per
mg/kg/day, based on renal tumors in treated male mice (EPA, 1998d).
The MCLG for BDCM is zero, based on its probable carcinogenicity and the use of linear
low-dose extrapolation in the cancer risk assessment. The MCLG of zero is protective for both
the carcinogenic and systemic effects of BDCM in children and adults. In addition, there is not
sufficient evidence indicating whether children are more sensitive to the toxic effects of BDCM
than are adults.
Dibromochloromethane (DBCM). Limited data are available for DBCM
reproductive/developmental toxicity. A prospective epidemiology study suggested that increasing
levels of DBCM were associated with significantly shorter menstrual cycles. Because the subjects
were also exposed to other contaminants and disinfection byproducts in the drinking water,
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establishment of a causal relationship between DBCM exposure and the observed effect is
difficult.
No developmental/reproductive toxicity was observed in a rat developmental study
(Ruddick et al., 1983) or in a rat short-term reproductive screening study (NTP, 1996).
However, there was marginal evidence for decreased fetal weight in a mouse two-generation
reproductive study (Borzelleca and Carchman, 1982).
DBCM has been shown to affect the nervous and immune systems, kidneys, and liver in
rats (NTP, 1985; Tobe et al., 1982). An RfD of 0.02 mg/kg/day was calculated based on
reported histologic lesions (fatty changes) in the liver in male rats after subchronic exposure to
DBCM (NTP, 1985).
Carcinogenicity studies have shown an increase in liver tumors in male and female mice
and no increase in tumors in rats (NTP, 1985). Under EPA's 1999 Draft Guidelines for
Carcinogen Risk Assessment, there is suggestive evidence of human carcinogenicity of DBCM,
but the data are not sufficient to assess human carcinogenic potential. Dose-response
assessment is not recommended under the EPA's Draft Guidelines for Carcinogen Risk
Assessment (EPA, 1999) for chemicals described by this descriptor. This descriptor is used when
the evidence from human or animal data is suggestive of carcinogenicity, but further studies
would be needed to determine human carcinogenic potential.
The MCLG for DBCM is 0.06 mg/L, based on the subchronic study in rats (NTP, 1985),
which was used to calculated the RfD of 0.02 mg/kg/day. Calculation of the MCLG took into
account a drinking water RSC of 80% and an additional safety factor of 10 to account for the
possible carcinogenicity of DBCM. EPA believes that the MCLG is protective for children's
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health, because developmental or reproductive effects have not been found to occur below the
level of the critical effect (liver toxicity) used as the basis for the MCLG, and because
development of the MCLG includes the standard UF of 10 for protection of sensitive populations.
There is, however, some uncertainty in this conclusion, because there is not sufficient evidence
from studies on the systemic effects of DBCM, or on the metabolism of the compound, to
determine whether children are more sensitive to the toxic effects of DBCM than are adults.
Bromoform. A prospective epidemiology study suggested that increasing levels of
bromoform were associated with significantly shorter menstrual cycles. Because the subjects
were also exposed to other contaminants and disinfection byproducts in the drinking water,
establishment of a causal relationship between bromoform exposure and the observed effect is
difficult.
Developmental toxicities have been reported in rats and mice treated with bromoform, and
these effects occurred at dose levels either below or equal to the dose that caused maternal
toxicity. In a rat developmental study (Ruddick et al., 1983), fetal skeletal anomalies were
observed at a dose of 100 mg/kg/day, while no maternal toxicity was observed at doses as high as
200 mg/kg/day. In a mouse two-generation reproductive study (NTP, 1989a), developmental
toxicities observed in F1 mice exposed to 200 mg/kg/day included hepatocellular degeneration,
decreased postnatal survival, and other signs of toxicity (increased relative liver and decreased
relative kidney weights), while the same dose also resulted in decreased body weight of pregnant
dams at delivery.
Systemic effects on the liver, kidney, nervous system, and thyroid have been noted in rats
and mice from bromoform exposure (NTP, 1989b; Tobe et al., 1982). Based on hepatocellular
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vacuolization in the liver of male rats in a subchronic study (NTP, 1989b), the Agency derived an
RfD of 0.02 mg/kg/day. Because the database for bromoform includes systemic toxicity studies in
two species, a two-generation study in mice, and a developmental toxicity study in rats, no
database UF was used in developing the RfD.
Carcinogenicity bioassays have reported an increase in large intestine tumors in male and
female rats, and no increase in male and female mice (NTP, 1989b). In a study in which
bromoform was administered by intraperitoneal injection in mice, the ratio of the number of lung
tumors per mouse in the mid-dose group was significantly elevated over controls (Theiss et al.,
1977). Based on the observation of a rare tumor type in both sexes of rats, supported by positive
genotoxicity findings, bromoform is considered likely to be carcinogenic in humans under EPA's
1999 Draft Guidelines for Carcinogen Risk Assessment, and a cancer oral slope factor of 4.5x 10"
3 per mg/kg/day was calculated (EPA, 2002c).
The MCLG for bromoform is zero, based on its probable carcinogenicity and the use of
linear low-dose extrapolation. EPA believes that the MCLG of zero is protective for both the
carcinogenic and systemic effects of bromoform in children and adults. In addition, there are no
studies indicating that children are more sensitive to the toxic effects of bromoform than are
adults.
Monochloroacetic acid (MCA). Developmental studies have shown mixed results. A
published abstract of a developmental study in rats reported the occurrence of malformations of
the cardiovascular system and maternal toxicity at the highest dose tested (Smith et al., 1990);
another study did not find developmental toxicity at a maternally toxic dose, but this study was
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limited by a small sample size and incomplete examinations (Johnson et al, 1998). No studies
were located on the reproductive toxicity of MCA.
Subchronic and chronic studies suggest that the primary targets for MCA-induced toxicity
are the spleen, heart and nasal epithelium (DeAngelo et al., 1997; NTP, 1992b). An RfD of 0.004
mg/kg/day was calculated based on a chronic rat drinking water study (DeAngelo et al., 1997) in
which the critical effect was increased spleen weight compared to controls. Due to the lack of
adequate developmental toxicity studies in two species and the lack of a multi-generation
reproductive study, a 3-fold UF for database deficiencies was used in the calculation of the RfD.
MCA has not been shown to be carcinogenic in two chronic rodent bioassays (DeAngelo
et al., 1997; NTP, 1992b). However, both studies suffer from methodological limitations that
limit the strength of the conclusions that can be drawn from them. Under the 1999 Draft
Guidelines for Carcinogen Risk Assessment (EPA, 1999), the data on MCA are inadequate for an
assessment of human carcinogenic potential.
The proposed MCA MCLG is 0.03 mg/L, based on the RfD of 0.004 mg/kg/day and
assuming a RSC of 20% (EPA, 2002a). EPA believes that this MCLG is protective for children
since the levels at which systemic toxicity have been seen are much lower (20x) than the levels at
which developmental effects have been observed. There is no evidence from studies on the
systemic effects of MCA that children are more sensitive to the toxic effects of MCA than are
adults. However, these data are limited. No data on potential metabolic differences between
children and adults for MCA were located.
Dichloroacetic acid (DCA). Developmental effects have been reported in two rat
developmental studies. Smith et al. (1992) reported that DCA treatment during rat pregnancy
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resulted in developmental toxicities including implantation loss, reduced number of live fetuses,
decreased fetal body weight and crown-rump length, cardiovascular, and soft tissue
malformations. Epstein et al. (1992) reported that acute high-dose treatments with DC A at
specific rat developmental stages can induce developmental toxicity (e.g., cardiac defects) in the
absence of maternal toxicity.
The liver, testes, and brain are the major target organs for DCA-induced systemic toxicity
in DCA-treated rats, mice, dogs, and humans (Cicmanec et al., 1991; DeAngelo et al., 1999; Katz
et al., 1981; Parrish et al., 1996; Stacpoole et al., 1998). The RfD of 0.004 was based on
testicular degenerative changes, liver vacuolization, and brain histopathology, observed in a
subchronic study in dogs (Cicmanec et al., 1991).
Studies in male mice (Anna et al., 1994; Bull et al., 1990; DeAngelo et al., 1991;
DeAngelo et al., 1999; Daniel et al., 1992; Ferreira-Gonzalez et al., 1995; Herren-Freund et al.,
1987), female mice (Pereira and Phelps, 1996; Pereira, 1996), and rats (DeAngelo et al., 1996)
have shown an increase in the incidence of liver tumors from DC A exposure. According to the
1999 Draft Carcinogen Risk Assessment Guidelines (EPA, 1999), DC A is likely to be
carcinogenic in humans, based on the weight of the evidence in animal bioassays and the lack of a
clear mechanistic understanding of the mode of action.
The MCLG for DC A is zero, based on its probable carcinogenicity and the use of linear
low-dose extrapolation. The Agency believes that this MCLG of zero is protective for both the
carcinogenic and systemic effects of DC A in children and adults. In addition, there is no evidence
from studies on the systemic effects of DC A that children other than those with metabolic defect
disorders are more sensitive to the toxic effects of DC A than are adults. However, children with
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hereditary tyrosinemia, glycogen storage disease or hyperoxaluria may be at increased risk for
DCA-induced toxicity, as are adults with the same defect disorders.
Trichloroacetic acid (TCA). Animal studies indicate that TCA can cause developmental
toxicity. A developmental study in rats (Johnson et al., 1998) reported increases in cardiac
malformations; this study was limited by a small sample size and incomplete examinations. In
another rat developmental study (Smith et al., 1989), TCA treatment during gestation resulted in
reduced fetal body weight and crown-rump length. In addition, soft-tissue malformations in the
cardiovascular system were increased in a dose-dependent manner. In both studies, maternal
toxicity was also observed at the doses that caused developmental effects. No reproductive
toxicity studies of TCA were identified.
TCA has been shown to affect the liver and kidney in rats and mice (DeAngelo et al.,
1997; Dees and Travis, 1994; Mather et al., 1990; Parrish et al., 1996; Acharya et al. 1995)
exposed to TCA in drinking water. The Agency based the TCA RfD on increased serum levels of
liver enzymes (indicating cell damage) and histopathological evidence of necrosis in the liver in a
2-year rat study (DeAngelo et al., 1997). A 10-fold database UF was used to account for
database deficiencies, including the lack of adequate developmental toxicity studies in two
species, the lack of a multi-generation reproductive study, and the lack of full histopathological
data in a second species. The calculated RfD is 0.03 mg/kg/day.
TCA has been shown to induce liver tumors in mice (Pereira, 1996) but not in rats
(DeAngelo et al., 1997). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment,
the data on TCA provide suggestive evidence of carcinogenicity, but the data are not sufficient
to assess human carcinogenic potential. Dose-response assessment is not recommended under
xxi
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the EPA's Draft Guidelines for Carcinogen Risk Assessment (EPA, 1999) for chemicals described
by this descriptor.
EPA has proposed an MCLG of 0.02 mg/L, based on an RfD of 0.03 mg/kg/day, a
drinking water RSC of 20%, and an additional safety factor of 10 to account for the possible
carcinogenicity of TCA (EPA, 2002a). This MCLG is expected to be protective of children, since
developmental and systemic toxicity appear to occur at similar doses, and there is no evidence
that children are more sensitive to the toxic effects of TCA than are adults. In addition,
uncertainties regarding the effect level for developmental toxicity have been taken into account
through the application of UFs.
Monobromoacetic acid (MBA). The toxicity data for MBA are very limited. There are
no peer-reviewed developmental studies, chronic toxicity, or carcinogenicity studies available. An
RfD has not been established for MBA, because there are no systemic toxicity studies of sufficient
duration. According to EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment (EPA,
1999), the data on MBA are inadequate for an assessment of human carcinogenic potential
(EPA, 2002b).
An MCLG has not been established for MBA. Data relevant to potential fetal sensitivity
are limited to a single developmental study reported in a published abstract in which maternal
effects and fetal effects (decreased live fetus size and increased incidence of soft-tissue
malformations) were noted in rats (Randall et al., 1991).
Bromochloroacetic acid (BCA). No long-term reproductive/developmental studies are
available for BCA. A short-term reproductive and developmental screening study reported
significant decreases in live fetuses/litter and total implants/litter; however, no treatment-related
xxii
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effects on sperm morphology or motility, mating or fertility were noted (NTP, 1998b). In
contrast, another short-term drinking water study reported significant impairment in sperm
motility, abnormal sperm morphology, altered spermiation, and reduced male fertility at
comparable drinking water concentrations (Klinefelter et al., 2002).
Oral toxicity studies have identified the kidney and liver as the target organs for BCA
toxicity (NTP, 1998b; Parrish et al., 1996). The data on BCA were insufficient for the derivation
of an RfD, because there are no systemic toxicity studies of sufficient duration.
There are no carcinogenic bioassays available for this compound (EPA, 2002b). Under
EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment (EPA, 1999), the data on BCA are
inadequate for an assessment of human carcinogenic potential.
Due to the limited data available on the noncancer or cancer effects of BCA, EPA has not
established an MCLG for this compound. Limited data suggest that decreased live fetuses/litter
and decreased total implants/litter occurred at doses comparable to the ones that caused general
toxicity in adults. The reproductive effects seen in male rats occurred at a lower dose, but no data
were identified on whether exposure of young males enhances their sensitivity. These data do not
support the hypothesis that fetuses or children are more sensitive than adults to the effects of
BCA..
Dibromoacetic acid (DBA). No peer-reviewed developmental studies are available, but
two published abstracts reported developmental effects, including increased prenatal mortality,
decreased pup weight, and hydronephrosis in gavage studies in mice (Narotsky et al., 1996;
Narotsky et al., 1997b). DBA has been shown to be spermatotoxic following high-dose single
exposures or repeated longer-term exposures (CCC, 2001; Linder et al., 1994a; Linder et al.,
xxiii
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1994b; Linder et al., 1995; Linder et al., 1997b), indicating that the male reproductive tract is a
target organ in both adult and developing rats. Based on available studies that evaluated male
reproductive toxicity in developing males (Klinefelter et al., 2000, 2001; Veeramachaneni et al.,
2000), it is unclear whether the developing organism is more sensitive to these effects. However,
a two-generation study found a lower incidence of males with reproductive toxicity in the F1 than
the parental generation, and thus developing males did not appear to be more sensitive than adults
in this study (CCC, 2001). In a standard one-generation drinking water study, Christian et al.
(2001b) found no DBA-related reproductive or developmental toxicity at drinking water
concentrations high enough to cause decreased water consumption due to poor palatability. In
contrast, recent results from Klinefelter et al. (2001) indicate that DBA causes pubertal delay
independent of effects on body weight.
DBA has been shown to affect the liver and immune system in drinking water studies in
mice (NTP, 1999; Parrish et al., 1996), and to cause neurobehavioral effects in a drinking water
study in rats (Phillips et al., 2002, published abstract). The data on DBA are insufficient for the
derivation of an RfD, because there are no systemic toxicity studies of sufficient duration.
No carcinogenicity bioassays are available for DBA (EPA, 2002b). In published abstracts,
So and Bull (1995) reported that DBA induces aberrant crypt foci in the colon of rats, and
Stauber et al. (1995) reported that DBA induces liver tumors in mice. Under EPA's 1999 Draft
Guidelines for Carcinogen Risk Assessment, there is suggestive evidence of carcinogenicity of
DBA, but the evidence is not sufficient to assess human carcinogenic potential. Dose-response
assessment is not recommended under the EPA's Draft Guidelines for Carcinogen Risk
Assessment (EPA, 1999) for chemicals for which this descriptor applies.
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Due to the limited data available on the noncancer and cancer effects of DBA, EPA is not
proposing an MCLG for the compound. The data are limited as to whether children may be more
sensitive than adults. A two-generation reproductive study (CCC, 2001) reported that the male
reproductive tract is a target organ in both adult and developing rats. However, the incidence of
affected animals was higher in the parental than the F1 generation, and thus developing males did
not appear to be more sensitive than adults. In a reproductive toxicity study (Christian et al.,
2001b), DBA readily crossed the placenta and distributed to the fetus, but it did not appear to
bioaccumulate. Developmental toxicity (prenatal mortality, decreased pup weight, and
hydronephrosis) was reported in gavage studies (Narotsky et al., 1996; 1997b), but not in the
drinking water reproductive toxicity studies (Christian et al., 2001b; CCC, 2001). It is unclear
whether this difference is due to differences in the route of dosing (gavage versus drinking water),
species sensitivity (mouse versus rat), or other factors. Regardless of the reason for the
difference, any developmental toxicity appears to occur at doses well above those causing male
reproductive toxicity. Additional reproductive studies are being conducted currently to evaluate
this effect in males.
Bromate. Limited data are available on the reproductive/developmental effects of
bromate. In a screening reproductive/developmental study in rats, a statistically significant
decrease in epididymal sperm density was observed (Wolf and Kaiser, 1996).
Acute oral poisoning of children with potassium bromate from accidental ingestion of hair
home permanent neutralizing solution have shown to result in central nervous system effects such
as sedation and lethargy, irreversible deafness and kidney effects (Benson, 1951; Gradus et al.,
1984; Lichtenberg et al., 1989; Lue et al., 1988; Mack, 1988; Parker and Barr, 1951; Quick et al.,
XXV
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1975; Warshaw et al., 1985; Watanabe et al., 1992). Subchronic and chronic animal studies
indicate that the kidney is the primary target organ for bromate toxicity following long-term
exposure (DeAngelo et al., 1998; Kurokawa et al., 1986b; Kurokawa et al., 1990; Nakano et al.,
1989). EPA has derived an RfD of 0.004 mg/kg/day for bromate, based on urothelial hyperplasia
in male rats in the DeAngelo et al. (1998) study (EPA, 2001b). A 3-fold database UF was used to
account for database deficiencies, including the lack of developmental toxicity studies in two
species and the lack of a multi-generation study.
A carcinogenicity study reported bromate exposure via drinking water to result in an
increase in tumors of the kidney, thyroid and tunical vaginalis in male rats and male mice
(DeAngelo et al., 1998). Under the Draft Guidelines for Carcinogen Risk Assessment (EPA,
1999), bromate is likely to be carcinogenic to humans by the oral route of exposure (EPA,
2001b).
The MCLG for bromate is zero, based on its probable human carcinogenicity and the use
of linear low-dose extrapolation. The Agency believes that the MCLG of zero is protective for
both the carcinogenic and systemic effects of bromate in children and adults. In addition, there
are no data suggesting that children are more sensitive than adults to bromate,
Chlorite/Chlorine Dioxide. Chlorite and chlorine dioxide are evaluated together in this
document because it is likely that the studies conducted with chlorite, the predominant
degradation product of chlorine dioxide, are relevant to characterizing the toxicity of chlorine
dioxide.
A number of studies conducted with chlorite and chlorine dioxide in rats, mice, and
rabbits have reported developmental/reproductive effects, including neurobehavioral effects,
xxvi
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increased number of resorbed and dead fetuses, and skeletal abnormalities (CMA, 1996; Couri et
al., 1982; Harrington et al., 1995b; Mobley et al., 1990; Suh et al., 1983; Moore et al., 1980;
Orme et al., 1985; Taylor and Pfohl, 1985; Toth et al., 1990). Lowered auditory startle amplitude
and altered liver weights in two generations of rats have been identified as the critical effects for
exposure to chlorite or chlorine dioxide (CMA, 1996).
Subchronic and chronic animal studies reported mixed results on the systemic toxicity of
chlorite or chlorine dioxide. Reported systemic effects included nasal and stomach lesions, spleen
alteration, adrenal weight change, and hematological effects. However, no firm conclusions can
be made regarding these observed systemic effects. An RfD for each of the two chemicals was
calculated to be 0.03 mg/kg/day (EPA, 2000c) based on the critical effects of lowered auditory
startle amplitude and decreased liver weight observed in the two-generation study in rats (CMA,
1996).
No chronic carcinogenicity bioassays are available for chlorine dioxide. Under EPA's
1999 Draft Guidelines for Carcinogen Risk Assessment, the data for chlorite and chlorine dioxide
human carcinogenicity are inadequate for an assessment of human carcinogenic potential.
An MCLG of 0.8 mg/L for chlorite and an MRDLG at the same level for chlorine dioxide
can be calculated from the RfD and using a drinking water RSC of 80%. This level is considered
to be protective of children because it is based on data from a two-generation reproduction study
that examined numerous developmental, reproductive and systemic endpoints. The results of the
CMA (1996) study are supported by the results of four other developmental studies that showed
similar effects at similar dose levels. In addition, a 10-fold factor was used in the derivation of the
xxvii
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RfD to account for human variability in response to the toxic effects of these chemicals, including
the response of sensitive individuals such as children.
Chlorine. Animal studies in rats and mice have not shown evidence of reproductive or
developmental effects from chlorine exposure (Carlton and Smith, 1985; Druckrey, 1968).
A 2-year study of chlorine in drinking water reported no systemic toxicity and no
carcinogenic effects in rats or mice (NTP, 1992a). The RfD of 0.1 mg/kg/day for chlorine was
calculated based on a free-standing NOAEL identified in a 2-year bioassay in female rats (NTP,
1992a).
Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, the data on chlorine
are inadequate for an assessment of human carcinogenic potential.
An MRDLG of 4 mg/L was calculated based on the RfD. EPA believes this level is
protective of children because the current RfD is based on a free standing NOAEL in a two-year
study in which chlorine dosing began when rats and mice were as young as 7 weeks old. In
addition, the UF used in the derivation of the RfD includes a 10-fold factor to account for human
variability in response to the toxic effects of these chemicals, including the response of sensitive
individuals such as children. In addition, aqueous chlorine appears to react so rapidly with sulfur-
containing amino acids in saliva that all free available chlorine is dissipated before water is
swallowed (EPA, 1994d). Therefore, the possibility of oral exposure to chlorine by fetuses,
infants and children is very limited.
Chloramine. The data on developmental effects of chloramine are limited to a rat
developmental study on monochloramine in drinking water (Abdel-Rahman et al., 1982).
-------
Monochloramine did not produce any significant changes in rat fetuses at any dose level. No
reproductive toxicity studies of chloramine were located.
A drinking water study in rats exposed for 12 months reported hematologic effects from
monochloramine exposure (Abdel-Rahman et al., 1984b); however, these effects were not
reported in a chronic study (NTP, 1992a), and the biological significance of the observed effects
is unclear. In a lifetime study of chloraminated water in rats (NTP, 1992a), there were no
biologically significant effects including body weight, organ weight, or histopathology at any dose
levels. The chloramine RfD of 0.1 mg/kg/day was calculated based on the absence of adverse
effects in the lifetime study in rats (NTP, 1992a).
No data were identified on the carcinogenic effects of chloramine in humans or animals.
Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, the data on chloramine are
inadequate for an assessment of human carcinogenic potential.
An MRDLG of 3 mg/L for chloramine (4 mg/L measured as total chlorine) was derived,
based on the lack of toxic effects, assuming a drinking water RSC of 80%. Data on
hematological effects of chloramine in infants and young animals are not available, but these
groups may be more sensitive than adults to hematological effects, because infants have a
transient deficiency of methemoglobin reductase, the enzyme that reduces methemoglobin to
hemoglobin. Nevertheless, EPA believes that the MRDLG for chloramine, which is based on a
free-standing NOAEL in a two-year study and includes the standard UF of 10 for protection of
sensitive populations, is protective of infants, although there is some uncertainty in this conclusion
in the absence of a quantitative analysis, in light of the possibility that newborns may have
increased sensitivity to the hematological effects of chloramine.
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MX (3-Chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone). A study in rats
showed no developmental toxicity at any MX dose tested, including a maternally toxic dose
(Huuskonen et al., 1997).
Subchronic rat studies have shown MX to have a strong local irritant effect on the
gastrointestinal tract (Vaittinen et al., 1995). An RfD has not been established for MX.
MX has been shown to be a carcinogen in rats, causing tumors of the thyroid, liver,
mammary gland, lung, adrenal gland, and pancreas (Komulainen et al., 1997). Under EPA's 1999
Draft Guidelines for Carcinogen Risk Assessment, MX is likely to be carcinogenic to humans
(EPA, 2000e).
EPA has not established an MCLG for MX because the Agency has not yet conducted a
full assessment of the systemic and carcinogenic effects of MX. There are insufficient data
available to evaluate whether children are more sensitive to the toxic effects of MX than are
adults.
Comparison of Toxicity Endpoints. Table ES-1 shows the MCLGs, MRDLGs, and the
toxicity endpoints used to set the MCLGs/MRDLGs for each of the disinfectants and disinfection
byproducts evaluated in this document.
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Table ES-1. Comparison of Toxicity Endpoints
Disinfectant/
Disinfection
Byproduct
Systemic Toxicity
(mg/kg/day)
Developmental Toxicity
(mg/kg/day)
Carcinogenicity
Based on 1999 Guidelines
MCLG/
MR I) IX,1
mg/L
NOAEL'
LOAEL3
NOAEL
LOAEL
Chloroform
None
(BMDL10 1.0)
12.9
35-50
126
likely to be carcinogenic to humans
under high exposure conditions0
0.07 (sys tox)
Bromodichloro-
methane (BDCM)
None
(BMDL10 0.8)
6d
25
50
likely to be carcinogenic to humans
0 (Ca)
Dibromochloro-
methane (DBCM)
None
(BMDL10 1.6)
28d
17
171
suggestive evidence of human
carcinogenicity, but the data are
not sufficient to assess human
carcinogenic potential.
0.06 (sys tox & Ca)
Bromoform
None
(BMDL10 2.6)
36d
50
100
likely to be carcinogenic to humans
0 (Ca)
Monochloroacetic
acid (MCA)
ND
3.5
70e
140
data are inadequate for an
assessment of human carcinogenic
potential
0.03 (sys tox)
Dichloroacetic
acid (DCA)
ND
12.5
14
140
likely to be carcinogenic to humans
0 (Ca)
Trichloroacetic
acid (TCA)
32.5
364
ND
291
suggestive evidence of
carcinogenicity, but the data are
not sufficient to assess human
carcinogenic potential
0.02 (sys tox)
Monobromoacetic
acid (MBA)
ND
ND
50e
100
data are inadequate for an
assessment of human carcinogenic
potential
NDf
Bromochloroacetic
acid (BCA)
15B
39
19
50
data are inadequate for an
assessment of human carcinogenic
potential
NDf
xxxi
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Disinfectant/
Disinfection
Byproduct
Systemic Toxicity
(mg/kg/day)
Developmental Toxicity
(mg/kg/day)
Carcinogenicity
Based on 1999 Guidelines
MCLG/
MR I) IX,1
mg/L
NOAEL3
LOAEL3
NOAEL
LOAEL
Dibromoacetic
acid (DBA)
2
10
50d
100
suggestive evidence of
carcinogenicity, but the data are
not sufficient to assess human
carcinogenic potential
NDf
Bromate
1.1
6.1
7.7
22
likely to be carcinogenic to humans
0 (Ca)
MX
ND
ND
ND
ND
likely to be carcinogenic to humans
NDf
Chlorite
ND
ND
3
6
data are inadequate for an
assessment of human carcinogenic
potential
0.8 (devel tox)
Chlorine dioxide
ND
ND
3
6
data are inadequate for an
assessment of human carcinogenic
potential
0.8 (devel tox)
Chlorine
14
ND
5h
ND
data are inadequate for an
assessment of human carcinogenic
potential
4 (sys tox)
Chloramine
9.5
ND
10
ND
data are inadequate for an
assessment of human carcinogenic
potential
3 (sys tox)1
aNOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse effect level; ND = not determined; BMDL10 = 95% lower confidence limit on the benchmark dose for 10% extra risk.
Where appropriate, duration-adjusted values shown
bMCLG = Maximum contaminant level goal; (Ca) = basis for MCLG is carcinogenic effects; (sys tox) = basis for MCLG is systemic toxic effects; (devel tox) = basis for MCLG is developmental effects.
cLikely to be carcinogenic to humans under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues, but not likely to be carcinogenic to humans at a dose
level that do not cause these effects.
dNo NOAEL was identified for the critical effect for BDCM or DBCM; the NOAEL for bromoform was 18.
eAdverse effect levels reported in a published abstract only, and thus should be considered preliminary,
insufficient data for derivation of a MCLG.
8Adverse effect level for systemic effects in a short-term study; no subchronic or chronic studies were available.
hHighest dose tested.
14mg/L measured as total chlorine
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PLAIN LANGUAGE SUMMARY OF EVALUATION OF CHILDREN'S RISK
It is the duty of EPA's Office of Water to make sure that the people living in the U.S.
have clean water and safe drinking water. To make the drinking water safe to drink, disinfectants
(such as chlorine and chlorine dioxide) are added to water to kill bacteria and other organisms
that can cause waterborne diseases in people of all ages. Waterborne disease can produce
illnesses with symptoms such as diarrhea, nausea, and stomach cramps. Disinfectants were first
added to drinking water many years ago, and their addition is today considered to be the most
successful public health program in U.S. history.
The Safe Drinking Water Act authorizes EPA to set national health-based standards for
drinking water to protect against both naturally-occurring and man-made contaminants that may
be found in drinking water and that may hurt people's health. U.S. EPA, states, and water
systems then work together to make sure that these standards are met. When EPA sets these
regulations, it considers the effects of the chemicals on many different groups within the general
population, including children, pregnant women, older people, and people with serious illnesses.
The disinfectants that kill bacteria and other organisms react with naturally-occurring material in
the water to form other chemicals called disinfection byproducts (DBPs), and the disinfectants and
disinfection byproducts (D/DBPs) may present the potential for health effects. Therefore, the
highest health protection requires balancing risks so that organisms that cause disease are
controlled, while also having a low risk from D/DBPs. The purpose of this document is to
examine whether children may be particularly sensitive to the health effects of drinking water
disinfectants and to D/DBPs.
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Studies in both animals and in people drinking chlorinated water with sufficiently high
levels of DBPs have shown that DBPs may present a health risk to people drinking the water.
Some of these chemicals have been shown to cause cancer in laboratory animals, and to have
effects on the developing fetus or on the liver and kidney in laboratory animals. Because it is very
hard to accurately measure exposure to DBPs, there is a lot of uncertainty about how much DBP
exposure may result in effects on fetuses or hurt one's ability to have children. The observed
relationship is also rather weak. Some data suggest that one specific class of DBPs, the
trihalomethanes, may cause the developmental effects, but it is not clear if this association is real,
or related to the methods used to measure an association. Some studies with people have
suggested that drinking water containing high levels of some of the DBPs may also increase a
person's risk of developing certain types of cancer. However, there are important differences in
the results from different studies, and it is not certain that the risk of any cancer is higher in people
drinking chlorinated water. There is very little information about whether children are more
sensitive than adults to any cancer-causing effects of DBPs.
EPA is using the health evaluations presented in this document for D/DBPs to set
Maximum Contaminant Level Goals (MCLGs) for DBPs or Maximum Residual Disinfectant
Level Goals (MRDLGs) for disinfectants. MRDLGs (and MCLGs) are nonenforceable health
goals based only on health effects and exposure information, and are established at the level at
which no known or anticipated adverse effects on the health of persons occur and which allows an
adequate margin of safety. Therefore, MRDLGs do not reflect the improvement to health
resulting from the disinfectant killing microbes and preventing disease. After EPA sets the
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MCLGs/MRDLGs, it then sets the enforceable drinking water regulations for the chemicals
(Maximum Contaminant Levels - MCLs), as close to the goal levels as is practical.
This document looks at the available health data for 13 DBPs and three disinfectants. For
four of the DBPs, EPA has determined that there are not enough data available on the potential
health effects to set an MCLG. For the other D/DBPs, EPA has determined that the MCLG or
MRDLG for each of these chemicals is protective of children. For some chemicals, this is because
the MCLG or MRDLG was set using a study that examined the effects of the chemical on the
developing fetus. For other chemicals, EPA determined that the harmful health effects were not
more likely to occur in children than in adults, or would not affect children at lower doses than
they would affect adults. This conclusion is based in part on the standard risk assessment
approach of using a protective factor to account for human variability, including potentially
sensitive populations such as children, although there is some uncertainty in this conclusion,
because data on general toxic effects in young animals are not available.
This document provides a summary of the existing or proposed MCLGs and MRDLGS
for each D/DBP. New or revised MCLGs are being proposed for chloroform, monochloroacetic
acid, and trichloroacetic acid.
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1. INTRODUCTION
1.1. RISK TO CHILDREN
EPA's Office of Water is charged with ensuring that the United States population has
clean water and safe drinking water. This mandate covers chemical, physical, and biological
pollutants. For drinking and other uses, water is often treated with chemical disinfectants to
prevent waterborne infectious disease. The disinfection of public drinking water supplies to
prevent waterborne disease is probably the most successful public health program in U.S. history.
The most common result of infection by waterborne pathogens is diarrheal disease. The diarrheal
diseases most commonly associated with waterborne infectious agents may have a more severe
outcome for children and debilitated adults; thus, the savings in lives from disinfecting water have
been largely in those populations. Chemical disinfection byproducts (DBPs) are formed from the
reaction of chlorine and other disinfectants with naturally occurring organic materials in source
water. These DBPs present the potential for health effects. Thus, to provide for maximum
human health protection it is necessary to balance risks from water pathogens and DBPs.
EPA's drinking water assessments and regulations have historically considered sensitive
subpopulations that may be at increased risk. Several initiatives in the mid to late 1990s mandated
explicit consideration of fetuses, infants and children as potentially sensitive subpopulations. In
1995, EPA established an agency-wide policy that calls for consistent and explicit consideration of
the risk to infants and children in all risk assessments and characterizations, as well as in
environmental and public health standards (Memorandum from the Office of the Administrator,
October 20, 1995). The Safe Drinking Water Act amendments of 1996 also stipulated that in
establishing maximum contaminant levels (MCLs) the Agency shall consider "the effect of such
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contaminants upon subgroups that comprise a meaningful portion of the general population (such
as infants, children, pregnant women, the elderly, individuals with a history of serious illness or
other subpopulations) that are identifiable as being at greater risk of adverse health effects due to
exposure to contaminants in drinking water than the general population." On April 21, 1997, the
President Clinton signed an Executive Order (13045) that federal health and safety standards must
include an evaluation of the potential risks to children in planned regulations. EPA's Office of
Water follows the above policy and order and has historically considered risks to sensitive
populations (including fetuses, infants and children) in establishing drinking water assessments,
advisories or other guidance and standards (Ware, 1989; EPA, 1991a).
There is a growing awareness that children, infants, and fetuses (like other developing
organisms) are not small adults, and may react to chemical exposures differently from adults.
These differences can result because the physiology of the developing body differs from that of
adults; children consume more food and water and inhale more air than adults in proportion to
their body mass; and the activity patterns of children, including crawling, mouthing, and outdoor
play, may result in greater exposure to environmental contaminants. From the physiological
perspective, the neurological, immunological, and digestive systems of children are still
developing, and thus can be more sensitive to chemical damage. These differences can lead to
impaired fetal or child development at doses that do not induce adverse effects in adults. In
addition, children undergo substantial hormonal changes as they enter sexual maturity. As
described in more detail below, chemical toxicokinetics (absorption, distribution, metabolism, and
excretion) can also differ between adults and children, particularly for children under one year of
age. Thus, this document considers hazards to children that may result from exposure during
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preconception, prenatal or postnatal development through sexual maturity. EPA must balance the
potential for greater risk to children from D/DBPs with the greater sensitivity of children to the
diarrhea resulting from infection with waterborne pathogens.
Epidemiology data show that cancer risk can also differ between children and adults. For
example, children were reported to be 2-fold more sensitive to the carcinogenicity of ionizing
radiation than adults (NRC, 1990). Recent experience has shown that diethylstilbestrol (DES)
treatment during pregnancy caused some female offspring to develop clear cell vaginal
carcinomas, but the same carcinomas did not appear at an increased rate in DES-exposed
mothers. In addition, it has been argued that developing organisms may be more sensitive than
adults to the effects of strong mutagens, because there is less time for DNA damage to be
repaired before the cell divides. A comparison of cancer bioassay results from 69 studies (EPA,
1996b) indicated that combined perinatal and adult exposure slightly increases the incidence of a
given type of tumor compared to the normal bioassay protocol, but it is not known if this reflects
the effect of an increased length of exposure or a heightened sensitivity of the young animal to the
carcinogenic effects of the chemical.
Age-dependent differences in toxicokinetics can also be an important determinant of
children's risk. Infants (and children, to a lesser degree) differ from adults in a number of factors
affecting absorption, distribution, metabolism, and excretion (reviewed by Scheuplein et al.,
2002). Gastric dynamics and enzyme levels differ from adult values, but generally reach adult
levels by the first year of life or earlier. Distribution can exhibit age-specific differences, due to
such factors as differences in total body fat and extracellular water. Age-related differences in
parameters affecting distribution are most apparent between adults and infants, although changes
3
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continue to occur during childhood, adolescence, and adulthood. Both chemical
biotransformation (Phase I) and conjugation (Phase II) metabolism systems are generally
immature at birth, and infants may have alternative metabolic pathways. Cytochrome P450
enzymes are the most common Phase I enzymes. These enzymes appear in human fetuses during
the first half of gestation, and enzyme activity levels at birth are about one-third the activity in
adults (Scheuplein et al., 2002). In infants, cytochrome P450 2E1 (CYP2E1), an isoenzyme of
cytochrome P450, rapidly increases after birth to a level similar to that in young/mature adults;
other isoenzymes of P450 have a low activity immediately after birth (EPA, 2000d). If a chemical
is converted to a less toxic form by CYP2E1, lower enzyme levels in infants would increase
toxicity. If a chemical is converted to a more toxic form by CYP2E1 metabolism (activation),
lower enzyme levels in infants would decrease toxicity. Conversely, children age 6 months up to
approximately 12 years can metabolize and eliminate some chemicals more rapidly than adults
(Renwick and Lazarus, 1998; Ginsberg et al., 2002). In these cases, increased enzyme activity in
young children for these chemicals would decrease toxicity if the chemical is converted to a less
toxic form and increase toxicity if the chemical is converted to a more toxic form. The potential
for enzyme induction (increases in enzyme levels following exposure to certain chemicals) may
also differ with age.
The expression of isoenzymes is also both species- and strain-dependent. In different
animal species, the development of isoenzymes follows different courses, producing different
patterns of age-related changes in metabolizing capability. For example cytochrome P450
enzymes are virtually absent in rat fetuses, but are present in humans in the latter half of
pregnancy (Scheuplein et al., 2002). After birth, the levels of at least one cytochrome P450
4
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enzyme (CYP2E1) increase rapidly in animals to adult levels (Song et al., 1986; Umeno et al.,
1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990). These interspecies differences need to
be taken into account in extrapolating from developing laboratory animals to children, and may
make this extrapolation more complex. Age-related changes in enzyme inducibility can also
occur.
In addition to differences in susceptibility engendered by physiological differences between
children and adults, differences in exposure may also contribute to differences in risk. For
example, children may incur greater exposure, because they ingest more food and fluids, and
inhale more air in proportion to their body weights than do adults. For example, the mean
drinking water ingestion rate for the U.S. population (all ages) is 16 mL/kg/day, while the mean
ingestion rate for babies younger than one year old is 46 mL/kg/day and for children aged 1-10 it
is 19 mL/kg/day. Thus, for babies, the drinking water ingestion rate is estimated to be three to
four times higher than for the population as a whole (EPA, 2000b). Similarly, the breathing rate
for children is higher than that for adults. A mean estimate of the breathing rate for children aged
1-2 years is 6.8 m3/day, corresponding to a mean body weight of 12 kg (EPA, 2000a). By
contrast, the mean inhalation rate is 11.3 m3/day for women and 15.2 m3/day for men, and the
mean adult body weight is 71.8 kg (EPA, 1997). This means that, for a chemical that is absorbed
from the respiratory tract and systemically distributed, a given concentration in air would result in
a higher dose on a mg/kg/day basis for children. Dermal exposure would also be higher, since
children have a higher surface area per unit body weight, as compared to adults. For example, the
average surface area:body weight ratio for children aged 0-2 is 0.064, while the average ratio for
adults is 0.028 (EPA, 1997). This means that children would have a higher exposure than adults
5
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to water contaminants from bathing, showering, or swimming. Exposure levels can also differ
due to different activity patterns, the increased hand-to-mouth activity of infants being a
commonly cited example. In summary, children may be more sensitive to the toxic effects of
certain environmental contaminants than are adults, due to their developing organ systems,
different metabolic rates and processes, higher surface area per unit weight, and greater exposure
potential. This document addresses the potential greater sensitivity of children, infants, and
fetuses to the toxic effects of chemicals, due to the potential for toxicokinetic and toxicodynamic
differences between adult and developing organisms.
1.2. RISK ASSESSMENT METHODS
Risk assessment is a process by which judgments are made about an agent's potential to
cause harm to humans. Risk assessment of chemicals follows the process developed by the
National Academy of Sciences/National Research Council (NAS/NRC). This process is based on
analysis of scientific data to determine the likelihood, nature and magnitude of harm to public
health associated with exposure to environmental agents (NRC, 1983, 1994). The NAS paradigm
defines four steps:
• Hazard Assessment— The process of determining whether exposure to an agent
can cause an increase in the incidence of a particular adverse health effect (e.g.,
cancer, birth defect) and whether the adverse health effect is likely to occur in
humans.
• Dose-Response Assessment—A determination of the relationship between the
magnitude of an administered, applied, or internal dose and a specific biological
6
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response. Response can be expressed as measured or observed incidence, percent
response in groups of subjects (or populations), or as the probability of occurrence
within a population
• Exposure Assessment— An identification and evaluation of the human population
exposed to a toxic agent, describing its composition and size, as well as the type,
magnitude, frequency, route and duration of exposure
• Risk Characterization—Summarizes and integrates the scientific findings of the
hazard, dose-response and exposure assessments to determine the potential human
risk, and discusses uncertainties and assumptions.
To evaluate the special toxicity of disinfectants and disinfection byproducts (D/DBPs) for
fetuses, infants, and children, the following three types of toxicity studies were evaluated:
1. Developmental and reproductive toxicity including both prenatal and postnatal
exposures and effects
2. Systemic toxicity
3. Carcinogenicity
These study types are evaluated in the context of EPA's guidelines for developmental
toxicity risk assessment (EPA, 1991b), reproductive toxicity risk assessment (EPA, 1996c), and
carcinogen risk assessment (EPA, 1999).
Developmental toxicity is defined as the occurrence of adverse effects in the developing
organism that may result from exposure before conception, during prenatal development, or
postnatally to the time of sexual maturation. Adverse effects may be detected at any point in the
life span of the organism (i.e., in the developing organism, neonate, adolescent, or even in the
7
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elderly as a late-age-onset disorder). There are a number of developmental abnormalities of
concern: spontaneous abortions, stillbirths, premature mortality, reduced birth weight, reduced
crown-rump length, malformations (overt teratogenicity, such as spinal bifida, no eyes, cleft
palate, hydrocephaly, or anencephaly), delayed skeletal ossification, mental retardation, and
sensory loss, as well as other adverse functional and physical effects. Developmental
abnormalities from all causes are extremely common and present an enormous burden for society.
Reproductive toxicity is defined as the occurrence of biologically adverse effects on the
reproductive systems of females or males that may result from exposure to environmental agents.
This may be expressed as alterations to the female or male reproductive organs, the related
endocrine system, or pregnancy outcomes (EPA, 1996c). Some examples include adverse effects
on fertility, gestation, female reproductive cycle normality, the onset of puberty, or premature
reproductive senescence. In males, reproductive toxicity may be manifest as adverse effects on
male reproductive organ weight and morphology, sexual behavior, sperm count, sperm
morphology, and sperm activity. Male reproductive effects are indirectly related to children's
health risk due to the potential for male fertility problems to affect the ability of a woman to
conceive and the subsequent health of the fetus and child.
In the quantification of noncarcinogenic effects of oral exposure, a reference dose (RfD) is
most often calculated. The RfD is an estimate (with uncertainty spanning an order of magnitude)
of a daily exposure to the human population (including sensitive subgroups) that is likely to be
without an appreciable risk of deleterious health effects during a lifetime. The RfD is derived
from a no-observed-adverse-effect level (NOAEL), or lowest-observed-adverse-effect level
(LOAEL), identified from a chronic or subchronic study, divided by UF(s) (UF). When the data
8
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support it, a benchmark dose2 (BMD) (technically, the 95% lower confidence limit on the BMD,
termed the BMDL) may be calculated by applying an appropriate mathematical curve-fitting
procedure. The BMDL can then be used as a point of departure for calculation of the RfD. The
RfD is calculated as follows:
RfD __ (NOAEL or LOAEL or BMDL) __ _ mg/kg/day
Uncertainty Factor(s)
Selection of the UF to be employed in the calculation of the RfD is based on professional
judgment, which considers the entire database of toxicologic effects for the chemical. UFs are
applied as described in the box on the next page.
The Agency is increasingly moving towards incorporating toxicokinetic and
toxicodynamic data in the development of UFs, moving away from the use of default UFs. There
is a continuum in the application of UFs, ranging from default, through categorical, to fully data-
derived values based on chemical-specific data. The International Programme on Chemical Safety
(IPCS) has developed guidelines for the use of empirical data in the development of Chemical-
Specific Adjustment Factors (CSAFs) for interspecies differences and human variability (IPCS,
2001). In the framework adopted by the IPCS, the default UFs of 10 for interspecies differences
and 10 for intraspecies differences are each divided into subfactors for toxicodynamics and
toxicokinetics. When data are available, the default factors can be replaced with categorical
uncertainty factors, also called Chemical-Specific Adjustment Factors (CSAFs). These guidelines
2A11 benchmark doses reported in this document reflect the lower confidence limit on the dose that
would result in a 10% extra risk, i.e., a benchmark response (BMR) of 10%, notated as the BMDL10. The
benchmark doses (BMDs) reported here reflect the central tendency estimate, while the BMDL values
reflect the 95% lower confidence limit on the dose.
9
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allow data to be used quantitatively in the development of UFs, rather than relying solely on
professional judgement and the standard factors of 1, 3, and 10 noted in the box.
The potential for greater sensitivity of children, infants, and fetuses is also taken into
account in several aspects of the assessment. For example, this issue is considered in the choice
of UF for the adequacy of the database; part of the consideration for the database UF is whether
reproductive and developmental toxicity studies are available. With the greater attention being
paid to children's risk, the choice of the database UF also considers whether there is any reason
for concern that developing organisms may be more sensitive to systemic toxicity caused by the
chemical. In addition, the 10-fold factor for human variability, includes variability in the response
of sensitive individuals such as children.
Cancer assessment includes both qualitative and quantitative components. The qualitative
assessment consists of an evaluation of the weight-of-evidence of the carcinogenic risk from the
chemical. EPA's 1986 Guidelines for Carcinogen Risk Assessment evaluated chemicals using a
letter classification system (A-E) indicating the weight-of-evidence of carcinogenic risk (EPA,
1986). This classification was based primarily on animal and human studies showing an increase
in tumors. In 1996, EPA proposed revisions to the 1986 Guidelines (EPA, 1996a). Further draft
revisions to the Guidelines were released in 1999 (EPA, 1999). EPA has adopted the policy that,
until final guidelines are issued, the 1999 draft revised guidelines will serve as EPA's interim
guidance to EPA risk assessors conducting cancer risk assessments (EPA, 200 lg). Rather than
relying exclusively on tumor findings, the new draft Guidelines (EPA, 1999) include an expanded
weight-of-evidence approach that emphasizes understanding mode of action (MO A), conditions
of expression of carcinogenicity (e.g., route and magnitude of exposure) and consideration of all
10
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other relevant data. A narrative with descriptors replaced the letter classifications used in 1986
guidelines. The descriptors in the draft Guidelines (EPA, 1999) are: carcinogenic to humans;
likely to be carcinogenic to humans; suggestive evidence of carcinogenicity but not sufficient to
assess human carcinogenic potential; data are inadequate for an assessment of human
carcinogenic potential; and not likely to be carcinogenic to humans.
Under the 1999 Draft Guidelines, the preferred approach for dose-response assessment is
to use a biologically-based dose response (BBDR) model for extrapolating from animal to human
doses, and for extrapolating to low doses, if the necessary data for the parameters used in such
models are available. In the absence of such data, the default approach for calculating the human
equivalent dose is to scale the daily applied lifetime oral dose in proportion to body weight raised
to the 3/4 power (BW3 4). In the absence of a BBDR, dose-response assessment is done in two
steps. First, appropriate models are fit to data in the empirical range of observation to determine
a point of departure (POD). A standard POD is the effective dose corresponding to the lower 95
percent limit on a dose associated with 10 percent increased tumor or relevant nontumor response
(LEDio). In the second step, extrapolation is done below the observable range, to doses that are
more characteristic of human exposures. One of several default approaches (linear, nonlinear or
both) can be used for low-dose extrapolation, depending on the cancer MO A. The linear
approach is used when there is an absence of sufficient tumor MOA information, the chemical has
direct DNA mutagenic reactivity or other indications of DNA effects that are consistent with
linearity, or the dose-response relationship is expected to be linear. The nonlinear approach is
used when a tumor MOA supports nonlinearity, and the chemical does not demonstrate mutagenic
effects consistent with linearity, or the chemical has some indication of mutagenic activity, but it is
11
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judged not to play a significant role in tumor causation. If the MOA indicates both linear and
non-linear dose responses, both linear and nonlinear approaches are used. The 1999 Draft
Guidelines also provide additional guidance for the examination of the risk to children and other
sensitive populations. Based on linear extrapolation from the POD, the cancer slope factor, in
units of (mg/kg/day)"1 can be calculated. The slope factor can be converted to a drinking water
unit risk, in units of (ug/LH by dividing by 70 kg, multiplying by 2 L/day, and adjusting the units.
12
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Uncertainty Factors Used in RfD Calculations
Standard Uncertainty Factors (UFs)
Use a 1, 3, or 10-fold factor when extrapolating from valid experimental results from studies
using prolonged exposure to average healthy humans. This factor is intended to account for
the variation in sensitivity among the members of the human population. The intermediate
factor of 3 represents approximately '/S log10 unit, (i.e., the square root of 10).
Use an additional factor of 1, 3, or 10 when extrapolating from valid results of long-term
studies on experimental animals when results of studies of human exposure are not available or
are inadequate. This factor is intended to account for the uncertainty in extrapolating animal
data to risks for humans.
Use an additional factor of 1, 3, or 10 when extrapolating from less than chronic results on
experimental animals when there are no useful long-term human data. This factor is intended
to account for the uncertainty in extrapolating from less than chronic NOAELs to chronic
NOAELs.
Use an additional factor of 1, 3, or 10 when deriving an RfD from a LOAEL instead of a
NOAEL. This factor is intended to account for the uncertainty in extrapolating from LOAELs
to NOAELs.
Use an additional 3- or 10-fold factor when deriving an RfD from an "incomplete" data base.
The minimum database for a high confidence RfD is: two systemic toxicity bioassays in
different species, one two-generation reproductive study and two developmental toxicity
studies in different species. This factor is meant to account for the inability of any single type
of study to consider all toxic endpoints. The intermediate factor of 3 is often used when there
is a single data gap exclusive of chronic data.
The maximum composite UF for any given database is 3,000. Databases weaker than this are
judged too uncertain to estimate an RfD.
Note: With each UF assignment, it is recognized that professional scientific judgment must be
used.
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1.3 DETERMINING RISK TO CHILDREN
In developing the Stage 2 D/DBP Rule, risks to sensitive subpopulations including fetuses
and children were taken into account in the assessments of D/DBPs. To determine whether
fetuses and children are more sensitive than adults, the following questions were considered,
1. Is there information that shows that the D/DBP causes effects in the developing
fetus or harms a woman's ability to become pregnant and bear children? If it
causes these effects, do these effects occur at lower doses than those that cause
other types of effects?
2. If the D/DBP causes a health effect other than cancer, such as an effect on the liver
or kidney, are children affected at lower doses than are adults?
3. If the D/DBP causes cancer, are children more likely to be affected by a given dose
than are adults?
The ultimate goal of these questions is to determine whether the MCLG is protective of
any putative special risk to children, regardless of whether the MCLG is based on developmental
toxicity, systemic toxicity, or cancer effects.
The first question is addressed by directly comparing the dose-response in developmental
and reproductive toxicity studies with that in systemic toxicity and cancer studies. If the most
sensitive endpoint for a chemical is a developmental or reproductive effect, the RfD is based on
that endpoint. If adequate developmental or reproductive toxicity studies are not available, this
data gap is taken into account in the choice of the database UF.
14
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Ideally, to address questions 2 and 3 above, one would have data from a single study
comparing the sensitivity of young animals and adult animals exposed under the same conditions
and for the same duration. Such data are rarely available. In the absence of such a direct
comparison, one can obtain some relevant information on relative sensitivities by comparing the
NOAEL and LOAEL for effects in the first filial generation (F, pups) in a two-generation
reproductive study with the NOAEL and LOAEL in a subchronic study. This is because the two
study designs involve similar exposure durations, and the animals in the subchronic study are
exposed as young adults while the Fx pups in the reproductive study are exposed in utero through
young adulthood. Therefore, if everything else were the same between studies, differences in
toxicity could be attributed to age-related differences in sensitivity. This approach provides a very
crude comparison, since there may be differences between the subchronic and two-generation
studies, such as differences in dose spacing and in the strains used, complicating the comparison.
For most of the D/DBPs, appropriate data from a two-generation study were not available.
However, such a comparison was conducted for chloroform (see Section 2.2.1.3).
Since one rarely has adequate data to directly compare toxicity in adult and young
animals, and the comparison described for chloroform is rather crude, questions 2 and 3 are
usually addressed by considering all available data addressing the potential for age-related
differences in sensitivity, taking into account the factors noted in Section 1.1 that can lead to
increased sensitivity of children. Age-related differences in metabolism are identified where
possible, and these differences are evaluated in the context of whether the parent or a metabolite
is the toxic form. As noted earlier, interspecies differences in age-related metabolic capabilities
also need to be considered. Age-related differences in metabolism was considered for the
15
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trihalomethanes and the haloacetic acids, although a detailed analysis was possible only for the
trihalomethanes, because neither the mode of action nor the toxic forms are known for the
haloacetic acids. Where possible, potential age-related toxicodynamic differences are also
considered. Studies investigating toxicodynamic differences between young and adult animals are
rare. However, the consideration for chloramine noted that infants may be more sensitive than
adults to the hematological effects of this chemical.
An additional facet of these questions is consideration of whether differences between
adults and children are so significant that systemic toxicity in children would occur at a lower
dose than developmental toxicity, even though systemic toxicity in adults occurs at higher doses.
In other words, an additional question would be "does an RfD based on developmental effects as
the critical effect for the available database adequately protect children from systemic effects?"
This concern can be addressed by comparing toxicity in the two-generation and subchronic
studies, as described above. In the absence of such data, recent assessments paying special
attention to children's risk have begun to address this uncertainty using the database UF. For
older assessments, this remains an uncertainty.
The third question addresses the potential for children to be more sensitive than adults to
the carcinogenic effects of a chemical. To address this issue in general, one considers the
chemical's mode of action and metabolic pathway to determine whether a higher cancer risk may
be expected in children. Guidelines for quantitatively considering this issue are still in
development. However, such quantitative consideration is not needed to determine whether
MCLGs based on cancer are protective of children. MCLGs for genotoxic carcinogens are zero,
and so any MCLG for a genotoxic carcinogen would be protective of children. Chloroform is the
16
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only D/DBP that causes cancer, but has an MCLG other than 0, based on its mode of action. An
analysis of the available data on chloroform metabolism and its mode of indicates that children are
not expected to be at greater risk than adults from the carcinogenic effects of chloroform.
1.4. MAXIMUM CONTAMINANT LEVEL GOAL AND MAXIMUM RESIDUAL
DISINFECTANT LEVEL GOAL
The U.S. EPA is mandated to publish a maximum contaminant level goal (MCLG) and
promulgate a national primary drinking water regulation (NPDWR) establishing a maximum
contaminant level (MCL) if it determines that the contaminant (1) may have an adverse effect on
the health of persons; (2) is known to occur or there is a substantial likelihood that the
contaminant will occur in public water systems with a frequency and at levels of public health
concern; and (3) the Agency judges that regulation of the contaminant presents a meaningful
opportunity for health risk reduction for persons served by public water systems.
As defined in the Safe Drinking Water Act amendments of 1996, the MCLG is the level of
a contaminant in drinking water below which there is no known or expected risk to health.
MCLGs are nonenforceable health goals. They are set at concentration levels at which no known
or anticipated adverse effects on the health of persons occur, and which include an adequate
margin of safety. Establishment of an MCLG for a specific contaminant is based on the available
evidence of carcinogenicity or noncancer adverse health effects from drinking water exposure
using EPA's Guidelines for Risk Assessment. The Stage 1 Disinfectants/Disinfection Byproducts
(Stage 1 DBP) Rule at 63 FR 69390 provides a detailed discussion of the process for establishing
MCLGs. MCLGs can be based on an RfD for general systemic effects, on a quantitative estimate
17
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for developmental or reproductive effects, or on a consideration of a cancer quantitative
assessment. For carcinogenicity, when a linear low-dose extrapolation is done, the MCLG is set
at zero.
The maximum residual disinfectant level goal (MRDLG) concept was introduced in the
Stage 1 D/DBP Rule to reflect the fact that these substances have beneficial disinfection
properties. As with MCLGs, MRDLGs are established at the level at which no known or
anticipated adverse effects on the health of persons occur and which allows an adequate margin of
safety. MRDLGs are nonenforceable health goals based only on health effects and exposure
information and do not reflect the benefit of the addition of the chemical for control of waterborne
microbial contaminants.
The MCLGs and MRDLGs are used as the basis for setting the Maximum Contaminant
Levels (MCLs), which are the enforceable drinking water standards. MCLs are set as close to the
MCLGs and MRDLGs as feasible, taking costs, treatment technology, and other considerations
into account.
The MCLG or MRDLG for drinking water is calculated from the RfD for a 70 kg adult
consuming 2 L of water per day and also takes into consideration the relative source contribution
(RSC) from drinking water. For risk assessment the Agency views the use of 2 L per day adult
drinking water consumption as appropriate, because it represents the 84th to 90th percentile of
U.S. adult drinking water consumption (EPA, 2000b). The 90th percentile for community water
consumption of young people aged 11 to 19 in the United States is 1.5 L/day, and for children
younger than 1 year old it is less than 1 L/day, with a mean of 0.5 L/day or less (EPA, 2000b).
Although a conservative estimate of water intake for infants, expressed as the water intake:body
18
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weight ratio, is estimated at more than 3-fold the estimate for adults. The MCLG/MRDLG
calculated from the RfD is generally assumed to be protective of children. This is for several
reasons. The RfD is calculated to be protective of the human population including sensitive
subgroups; UFs used in deriving the RfD include a factor for human variability in response.
Second, the RfD is protective of lifetime exposure at a certain level. For most chemicals, the
severity of the effect increases with increasing exposure duration. This means that it might be safe
for children to be exposed for a less-than-lifetime duration to a chemical at doses that are higher
than the RfD (on a body weight basis). This second consideration indicates that an important part
of the consideration of children's risk is whether the effects begin to appear after only a short
period of exposure, or whether lifetime exposure is required for the effects to appear. The
critical effect for the RfDs for all of the D/DBPs were based on studies of developmental effects,
or on effects in subchronic or chronic studies; none of the RfDs were based on a short-term study.
The RfDs for three of the chemicals (dibromochloromethane, bromoform, and dichloroacetic
acid) were based on subchronic studies. This duration is long enough (roughly 1/10 of the
animal's life, corresponding to 7 years for humans) that a year of exposure at slightly more than
the RfD appears unlikely to result in the critical effect. For the RfDs of some of the D/DBPs
based on developmental effects, the dose to the fetus is determined by maternal water
consumption. For chlorite and chlorine dioxide, the critical effect was observed in both the Fx and
F2 generations, and the critical exposure window was not identified.
A final consideration regarding the protectiveness of the RfD is that the RfD is defined as
an estimate within "an order of magnitude." This is usually taken to mean that the "true" RfD
may be a factor of three higher or lower. Because of this imprecision in the estimate of the RfD,
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differences in dose within a factor of three are often considered to lie within the imprecision of the
method. Based on all of these considerations, the MCLG calculated from the RfD is generally
assumed to be protective of adverse effects over a lifetime exposure, and the Agency believes that
the use of 2 L to calculate the MCLG provides sufficient protection to fetuses and children.
The RSC is derived by application of the exposure decision tree approach published in
EPA's Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human
Health (EPA, 2000f). The RSC is the fraction of an individual's total exposure allocated to
drinking water. The Agency factors an RSC contribution into the MCLG to ensure that the
contribution from drinking water does not cause the total exposure of an individual to a
contaminant over a lifetime to exceed the contaminant's RfD. When data are sufficient, the
Agency uses those data as the basis for the RSC from drinking water. If insufficient exposure
data are available, a default assumption of a 20% drinking water contribution is used. The
maximum value used for an RSC is 80%, to allow for the possibility of other sources of exposure.
The Stage 2 D/DBP Rule proposes MCLGs for the disinfection byproducts chloroform (a
trihalomethane, or THM), and the haloacetic acids, monochloroacetic acid (MCA) and
trichloroacetic acid (TCA). For the other D/DBPs, the proposed Rule reflects MCLGs
established in the Stage 1 D/DBP Rule: bromodichloromethane (BDCM), dibromochloromethane
(DBCM), bromoform, dichloroacetic acid (DCA), bromate and chlorite. This document also
addresses the health effects of some additional DBPs: monobromoacetic acid (MBA),
bromochloroacetic acid (BCA), dibromoacetic acid (DBA) and 3-Chloro-4-(dichloromethyl)-5-
hydroxy-2(5H)-furanone (MX); however, the data are insufficient to develop MCLGs for these
20
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chemicals. The current and proposed MCLGs are shown in Table 1, and the MRDLGs for the
disinfectants chlorine, chloramine and chlorine dioxide are shown in Table 2.
Table 3 summarizes the toxicity endpoints for the various D/DBPs. BDCM, bromoform,
DC A, bromate, and MX are considered likely human carcinogens under the 1999 Draft
Carcinogenic Risk Assessment Guidelines. MCLGs of zero were selected after consideration of
the mode of action of these chemicals, except for MX, for which an MCLG has not yet been
determined. The MCLG/MRDLG values for chloroform, DBCM, MCA, TCA, chlorine, and
chloramine were based on systemic toxicity. For chlorine dioxide and chlorite, the
MCLG/MRDLGs are calculated on data from neurodevelopmental studies. No MCLGs were
derived for MX, MBA, BCA, and DBA, because the data are insufficient at this time. The
analysis of the children's risk in relation to each MCLB/MRDLG value is presented on a
chemical-by-chemical basis in the following sections.
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Table 1. Disinfection Byproducts and their MCLGs Considered in this Document
Disinfection Byproducts
Current MCLG
(mg/L)
Proposed MCLG
(mg/L)
Chloroform
0
0.07
Bromodichloromethane (BDCM)
0
0a
Dibromochloromethane (DBCM)
0.06
0.06a
Bromoform
0
0a
Monochloroacetic acid (MCA)
b
0.03
Dichloroacetic acid (DCA)
0
0a
Trichloroacetic acid (TCA)
0.3
0.02
Chlorite
0.8
0.8a
Bromate
0
0a
aNo (new) value proposed in the Stage 2 D/DBP rule.
bNo current MCLG, based on the Stage 1 D/DBP rule
Table 2. Disinfectants and their MRDLGs
Considered in this Document
Disinfectants
MR I) IX, (mg/L)
Chlorine
4 (as Cl2)
Chloramine
3 (4 mg/L as Cl2)
Chlorine dioxide
0.8 (as C102)
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Table 3. Comparison of Toxicity Endpoints
Disinfectant/
Disinfection
Byproduct
Systemic Toxicity
(mg/kg/day)
Developmental Toxicity
(mg/kg/day)
Carcinogenicity
MCLG/
MR I) IX,1
mg/L
NOAEL'
LOAEL3
NOAEL
LOAEL
Based on 1999 Guidelines
Chloroform
None
(BMDL10 1.0)
12.9
35-50
126
likely to be carcinogenic to humans
under high exposure conditions0
0.07 (sys tox)
Bromodichloro-
methane (BDCM)
None
(BMDL10 0.8)
6d
25
50
likely to be carcinogenic to humans
0 (Ca)
Dibromochloro-
methane (DBCM)
None
(BMDL10 1.6)
28d
17
171
suggestive evidence of human
carcinogenicity, but the data are
not sufficient to assess human
carcinogenic potential.
0.06 (sys tox & Ca)
Bromoform
None
(BMDL10 2.6)
36d
50
100
likely to be carcinogenic to humans
0 (Ca)
Monochloroacetic
acid (MCA)
ND
3.5
70e
140
data are inadequate for an
assessment of human carcinogenic
potential
0.03 (sys tox)
Dichloroacetic
acid (DCA)
ND
12.5
14
140
likely to be carcinogenic to humans
0 (Ca)
Trichloroacetic
acid (TCA)
32.5
364
ND
291
suggestive evidence of
carcinogenicity, but the data are
not sufficient to assess human
carcinogenic potential
0.02 (sys tox)
Monobromoacetic
acid (MBA)
ND
ND
50e
100
data are inadequate for an
assessment of human carcinogenic
potential
NDf
Bromochloroacetic
acid (BCA)
15g
39
19
50
data are inadequate for an
assessment of human carcinogenic
potential
NDf
23
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Disinfectant/
Disinfection
Byproduct
Systemic Toxicity
(mg/kg/day)
Developmental Toxicity
(mg/kg/day)
Carcinogenicity
MCLG/
MR I) IX,1
mg/L
NOAEL3
LOAEL3
NOAEL
LOAEL
Based on 1999 Guidelines
Dibromoacetic
acid (DBA)
2
10
50d
100
suggestive evidence of
carcinogenicity, but the data are
not sufficient to assess human
carcinogenic potential
NDf
Bromate
1.1
6.1
7.7
22
likely to be carcinogenic to humans
0 (Ca)
MX
ND
ND
ND
ND
likely to be carcinogenic to humans
NDf
Chlorite
ND
ND
3
6
data are inadequate for an
assessment of human carcinogenic
potential
0.8 (devel tox)
Chlorine dioxide
ND
ND
3
6
data are inadequate for an
assessment of human carcinogenic
potential
0.8 (devel tox)
Chlorine
14
ND
5h
ND
data are inadequate for an
assessment of human carcinogenic
potential
4 (sys tox)
Chloramine
9.5
ND
10
ND
data are inadequate for an
assessment of human carcinogenic
potential
3 (sys tox)1
aNOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse effect level; ND = not determined; BMDL10 = 95% lower confidence limit on the benchmark dose for 10% extra risk.
Where appropriate, duration-adjusted values shown
bMCLG = Maximum contaminant level goal; (Ca) = basis for MCLG is carcinogenic effects; (sys tox) = basis for MCLG is systemic toxic effects; (devel tox) = basis for MCLG is developmental effects.
cLikely to be carcinogenic to humans under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues, but not likely to be carcinogenic to humans at a dose
level that do not cause these effects.
dNo NOAEL was identified for the critical effect for BDCM or DBCM; the NOAEL for bromoform was 18.
eAdverse effect levels reported in a published abstract only, and thus should be considered preliminary,
insufficient data for derivation of a MCLG.
8Adverse effect level for systemic effects in a short-term study; no subchronic or chronic studies were available.
hHighest dose tested.
14mg/L measured as total chlorine
24
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2. ESTIMATES OF RISK TO CHILDREN FOR STAGE II DISINFECTANTS/
DISINFECTION BYPRODUCTS
A variety of different types of data are used for characterizing the risks to children from
D/DBPs. This chapter begins by reviewing the relevant available epidemiology data, focusing on
studies evaluating the potential developmental and reproductive effects from consumption of
chlorinated drinking water. The epidemiology data on cancer risk from consumption of
chlorinated drinking water are also briefly reviewed. This chapter then reviews the available data
for each D/DBP, beginning with information on metabolism when available, and addressing the
D/DBPs' developmental/reproductive effects, systemic effects, and carcinogenic potential. The
section for each D/DBP concludes with a discussion of the derivation of the MCLG/MRDLG for
that chemical, and an evaluation of children's risk relative to the MCLG/MRDLG.
2.1. CHLORINATED DRINKING WATER
This section considers epidemiology studies on exposure to chlorinated drinking water
rather than to individual D/DBPs. There are no reliable studies on the reproductive or
developmental toxicity of chlorinated drinking water in animals. While several reproductive and
developmental toxicity studies have been conducted with chlorinated drinking water, they either
used distilled water or chlorinated tap water (EPA, 1994d). Since the composition of D/DBPs in
drinking water depends on source water characteristics and the chlorination procedures used,
neither of these approaches may adequately represent the composition of chlorinated drinking
water. Animal studies are not available to address whether there are age-related differences in the
25
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systemic response to chlorinated drinking water. In addition, animal carcinogenicity studies of the
D/DBPs did not evaluate the effects of perinatal or in utero exposure.
2.1.1. Developmental/Reproductive Effects
The available epidemiologic studies on the developmental and reproductive effects from
exposure to D/DBPs in drinking water may be divided into two major groups: 1) qualitative
exposure assessment studies that examine associations between the source of the water supply or
disinfection method and the risk of adverse developmental effects; and 2) quantitative exposure
studies that associate adverse reproductive outcomes with reported concentrations of D/DBPs in
water supplies. Developmental/reproductive outcomes examined in both types of studies include
low birth weight, intrauterine growth retardation/small for gestational age, pre-term delivery,
spontaneous abortion, stillbirth, and birth defects (Reif et al., 2000).
Qualitative exposure studies have investigated a number of developmental/reproductive
outcomes. These studies typically do not allow the evaluation of a dose-response and are limited
because the effects cannot be attributed to a single contaminant or group of contaminants.
One study (Yang et al., 2000) examined the relationship between drinking water
disinfection and low birth weight, but found no statistically significant effects. The study authors
conducted a study in Taiwan of the association between chlorination of drinking water and low
birth weight. They compared the incidence of low birth weight in 14 cities in which chlorinated
water supplied over 90 percent of the residents with the rate in 14 matched cities that used non-
chlorinated water. They examined records on 18,025 births and found no association between
consumption of chlorinated drinking water and low birth weight (<2500 grams). They did
26
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observe an increase in pre-term deliveries (<37 weeks) in those cities using chlorinated water
(odds ratio [OR] 1.34, 95% confidence interval [CI] 1.15-1.56).
Several qualitative studies compared differences in birth outcomes, including birth weight,
in populations served with untreated drinking water and water treated using differing methods of
disinfection. Kallen and Robert (2000) evaluated differences in birth outcome in Sweden based
on approximately 74,300 control births, 15,400 births in areas using chlorine dioxide-treated
water and 24,700 births in areas served by water treated with sodium hypochlorite. Exposure
data were based on published reports of drinking water disinfection practices in the municipalities
at the time of the births (1985 - 1994). Two effects were significantly related to water chlorinated
with sodium hypochlorite: short body length (<43 cm, OR 1.97 95% CI 1.30-2.97) and small
head circumference (<31 cm, OR 1.46, 95% CI 1.07-1.98). In addition, increases were seen in
pre-term delivery (<32 weeks gestation, OR 1.22, 95% CI 1.00-1.48) and low birth weight
(<1500 g, OR 1.11, 95% CI 0.90-1.36); the increases were not statistically significant. These
effects were not observed more frequently for births in areas served by water treated with chlorine
dioxide. There was no difference between the exposed and unexposed groups for neonatal death,
neonatal jaundice, congenital malformations or childhood cancer.
Kanitz et al. (1996) examined 548 births during 1988-1989 in Italy from women in a
community exposed to filtered water disinfected with chlorine dioxide, sodium hypochlorite or
both and compared them with 128 births from women in an Italian community using untreated
water. Total trihalomethane levels were 8-16 ppm in the water treated with sodium hypochlorite,
and 1-3 ppb in the water treated with chlorine dioxide. Levels of chlorine dioxide in the water
immediately after treatment were less than 0.3 mg/L and chlorine residue was less than 0.4 mg/L.
27
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Infants born to women in the chlorine dioxide-disinfected community had smaller cranial
circumference (OR = 2.2, 95% CI = 1.4-3.9) and smaller body length (OR = 2.0, 95% CI = 1.2-
3.3) as compared to infants born to women in the untreated water community. Statistically non-
significant increases in low-birth-weight infants (<2,500 g) and preterm deliveries were reported
in women living in the chlorine dioxide-disinfected area. Mothers exposed to water treated with
chlorine dioxide, but not to sodium hypochlorite-treated water, had a significantly higher
frequency of newborns with neonatal jaundice. The authors concluded that infants of women who
consumed drinking water treated with chlorine dioxide or sodium hypochlorite during pregnancy
were at higher risk for neonatal jaundice, cranial circumference <35 cm, and body length <49.5
cm. However, potential confounding factors, such as the lack of information on the quantity of
water consumed during the pregnancy, exposure to other chemicals in the water, nutritional and
smoking habits and age distribution of the women, limit the conclusions that can be drawn from
this study (EPA, 2000c).
Other qualitative studies looked for effects of chlorination on early miscarriages
(spontaneous abortions), stillbirths or birth defects. In a case-control study based in the Boston
metropolitan area, Aschengrau et al. (1989) found an increased risk of early miscarriage for
surface water consumption versus ground water or mixed water consumption (OR 2.2, 95% CI
1.3-3.6). No significant difference was found for chlorinated versus chloraminated water supplies.
Aschengrau et al. (1993) conducted an additional case-control study of drinking water quality and
the occurrence of adverse pregnancy outcomes; this included 1,039 congenital anomalies, 77
stillbirths, and 55 neonatal deaths among women who delivered infants during August 1977
through March 1980 in Massachusetts. After adjustment for confounding, the frequency of
28
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stillbirths was increased (but not statistically significant) for women exposed to chlorinated
surface water (OR 2.6, 95% CI 0.9-7.5). In addition, the risk of major malformations was also
increased (OR 1.5, 95% CI 0.7-2.1), but the increase was not statistically significant. The
increase in major malformations was mainly due to a significantly increased risk of respiratory
tract defects (OR 3.2, 95% CI 1.1-9.5) and urinary tract defects (OR 4.1, 95% CI 1.2-14.1).
Magnus et al. (1999) conducted an ecologic study in Norway on the relationship between
the consumption of chlorinated drinking water and birth defects in the period 1993-1995. They
used data from 1994 on water quality and disinfection practice. Water color was used as an
indicator for natural organic matter content. Of 141,077 births in the study, 1.8% had birth
defects. A significant increase in urinary tract defects (OR 1.99, 95% CI 1.10-3.57) was seen
when the high color (high organic matter) plus chlorination group was compared with those with
low color in the drinking water plus no chlorination. There were nonsignificant increases in all
malformations (OR 1.14, 95% CI 0.99-1.31) and neural tube defects (OR 1.26, 95% CI 0.61-
2.62) in the same groups. In an additional analysis of this same population reported in a published
abstract (Jaakkola et al., 1999), the risk of low birth weight, intrauterine growth retardation, or
prematurity was not increased for individuals exposed to chlorinated water.
Several quantitative exposure studies examined the association between D/DBPs and pre-
term delivery, low birth weight and/or other adverse developmental outcomes. Savitz et al.
(1995) evaluated risks of miscarriage (n=261), pre-term delivery (n=413) or low birth weight
(n=301) for total trihalomethane (TTHM) exposure in a case-control study in North Carolina.
Exposure in this study was evaluated by linking the women's dates of pregnancy to the nearest
quarterly average TTHM level obtained from the water supplier. This was multiplied by the
29
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average number of glasses of water consumed per day, corresponding to the fourth week of
pregnancy for miscarriage cases and the 28th week of pregnancy for preterm delivery and low birth
weight cases. Women in the highest sextile (sixth) of exposures had an increased risk of
miscarriages (OR 2.8, 95% CI 1.2-6.1), but no such association was evident in the second highest
sextile (OR 0.2, 95% CI 0.0-0.5) or when the data were analyzed by exposure tertiles (thirds).
The risk for pre-term delivery was not increased, and the slight increase in risk for low birth
weight in the upper tertile was not statistically significant.
Gallagher et al. (1998) evaluated associations between estimated TTHM concentrations in
municipal water supplies and pre-term delivery, low birth weight, and term low birth weight in a
retrospective cohort study of 1,893 births. TTHM levels were modelled based on hydraulic
characteristics of the water system and monitored TTHM levels in the distribution system.
Exposures were assigned to study participants based on the estimated exposure during the
women's third trimester of pregnancy. The exposure categories for estimated TTHM
concentrations were <20 |ig/L (referent category), 21-40 |ig/L, 41-60 |ig/L and >61 |ig/L.
Women exposed to TTHM concentrations >61 |ig/L had an increased risk for low birth weight
(OR 2.1, 95% CI 1.0-4.8) and term low birth weight (OR 5.9, 95% CI 2.0-17.0), although the
number of cases of term low birth weight was small for the exposed mothers (n=6).
Nuckols et al. (1995) conducted a pilot study in Colorado evaluating the use of
geographic information system (GIS) technology for analyzing drinking water epidemiology data.
They investigated the relationship between water disinfection in two public water systems: one
using chloramination resulting in very low levels of trihalomethanes (THMs); and the other using
chlorination resulting in higher levels of THMs. Health outcome data, including birth weight and
30
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gestational age for live births in two Colorado counties in 1990, was obtained from birth
certificates. Exposure classification was based on the census block group of residence of the
mother at the time of delivery and actual monitoring data for THMs. Preliminary analyses
indicated no significant association between low birth weight and water disinfection. However,
the proportion of babies who were pre-term (<38 weeks of gestation) was significantly higher
(relative risk 0.54, 95% CI 0.33-0.88) in the chloraminated water district.
Two studies investigated early-term miscarriage risk factors. The first of these is a
quantitative exposure study that examined the potential association between early-term
miscarriage and exposure to THMs (Waller et al., 1998). The second is a qualitative study that
examined the potential association between early-term miscarriage and tap water consumption
(Swan et al., 1998). Both studies used the same group of 5,144 pregnant women living in three
areas of California. In the Waller et al. (1998) study, additional water quality information from
the women's drinking water utilities were obtained so that THM levels could be determined. The
Swan et al. (1998) study provided no quantitative measurements of THMs (or DBPs), and thus
provided no additional information on the risk from chlorination byproducts.
In the Waller et al. (1998) study, water utilities that served the population provided THM
measurements taken during the time period when participants were pregnant. The TTHM level in
a participant's home tap water was estimated by averaging water distribution system TTHM
measurements taken during a participant's first 3 months of pregnancy; a similar approach was
followed for estimating concentrations of individual THMs. Women who drank five or more
glasses per day of cold home tap water containing at least 75 |ig/L of TTHM were considered to
be in the high TTHM group. Incidence of early-term miscarriage in this group was 15.7%,
31
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compared with an incidence of 9.5% among women with low TTHM exposure (drinking fewer
than five glasses per day of cold home tap water or drinking any amount of tap water containing
less than 75 |ig/L of TTHM). An adjusted odds ratio for early-term miscarriage of 1.8 (95% CI
1.1-3.0) was determined. Only high BDCM exposure was associated with early-term miscarriage
(adjusted OR = 2.0, 95% CI = 1.2-3.5). This was defined as drinking five or more glasses per
day of cold home tap water containing >18 |ig/L BDCM. After adjustment for exposure to other
THMs, the adjusted odds ratio for early-term miscarriage was 3.0 (95% CI 1.4-6.6). The authors
noted several potential limitations of these data, including misclassification of exposure, the fact
that concentration levels for most subjects were based on test results from a single day, and that
THM exposure from sources other than ingestion could not be fully characterized. In a follow-up
study, Waller et al. (2001) reanalyzed the exposure data from this study using several different
methods. One goal of the reanalysis was to reduce exposure misclassification. The study authors
reported a positive dose-response relationship between spontaneous abortion rate and an
exposure metric incorporating total trihalomethanes and personal ingestion. However, the
authors noted that it was not possible to determine whether the reanalysis actually reduced
exposure misclassification. This reanalysis did not address dose-response relationships between
individual trihalomethanes and occurrence of spontaneous abortion.
Additional quantitative exposure studies have investigated the association between water
chlorination and birth defects, stillbirths and other adverse developmental effects. Bove et al.
(1992) conducted a cross-sectional study for four northern New Jersey counties to evaluate
associations between total trihalomethane (TTHM) concentrations in water supplies and reported
birth weight, fetal deaths and birth defects. The study population consisted of 594 cases of
32
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stillbirths and 669 cases of birth defects. Birthweight, fetal deaths and birth defects were
evaluated by linking quarterly TTHM measurements from the water utilities with the mother's
address at the birth of the child. Exposures greater than 80 |ig/L TTHM were significantly
associated with term low birth weight (OR 1.29, 95% CI 1.08-1.5), and small for gestational age
(OR 1.14, 95% CI 1.04-1.3), although the association was weak. No association was reported
for very low birth weight, stillbirth or pre-term birth. Significantly elevated odds ratios were
reported for birth defects, including all defects (OR 1.53, 95% CI 1.14-2.1), central nervous
system defects (OR 2.52, 95% CI 1.40-4.5), and neural tube defects (OR 2.98, 95% CI 1.25-7.1).
The odds ratios for oral cleft defects (OR 1.74, 95% CI 0.88-3.4) and major cardiac defects (OR
1.84, 95% CI 0.95-3.6) were also elevated, but the increases were not statistically significant.
This same data set was reported in a published manuscript that included analysis using an
exposure level of 100 |ig/L TTHM as the criterion to separate subjects into high and low
exposure groups (Bove et al., 1995). The association between TTHM exposure and term low
birth weight (OR 1.42, 50% CI 1.22-1.65)3 and small for gestational age (OR 1.50, 90% CI 1.19
- 1.86) was significant, and the strength of the association increased with the higher criterial
exposure level. A mean decrease in term birthweight of 70.4 grams was associated with TTHM
exposures greater than 100 //g/L.
Klotz and Pyrch (1999) followed up on the work of Bove and colleagues (Bove et al.,
1992; 1995) and examined the potential association between neural tube defects and drinking
water contaminants including trihalomethanes, haloacetonitriles and haloacetic acids. In this
population-based case-control study, 112 births with neural tube defects reported to New Jersey's
3This was the only confidence interval reported by the authors for this endpoint.
33
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Birth Defects Registry in 1993 and 1994 were matched against control births chosen randomly
from across the state. Birth certificates were examined for all subjects, as were drinking water
data corresponding to the mother's residence in early pregnancy. The drinking water data was
obtained by analyzing tap water for haloacetic acids, haloacetonitriles, THMs, total chlorine and
free chlorine, four months after the mother's due date. This sampling was intended to correspond
to the critical period (one year earlier) of neural tube development in the fetus (i.e., taking into
account seasonal variation in the concentrations of these chemicals in drinking water). In
addition, TTHM exposure during the gestational period was estimated based on records from the
water utility. The authors reported elevated odds ratios, generally between 1.5 and 2.1, for the
association of neural tube defects with TTHMs. The only statistically significant results were seen
when the analysis was isolated to those subjects with the highest THM exposures (greater than 40
ppb) and was limited to those subjects with only neural tube defects and no other malformations
(OR 2.1, 95% CI 1.1-4.0).
Dodds and King (2001) conducted a retrospective cohort study of the relationship
between BDCM exposure and birth defects among 49,842 residents of Nova Scotia, Canada
between 1988 and 1995. Exposure was estimated from routine water monitoring samples
collected from within the water distribution system. The birth defects examined had previously
been reported in other epidemiological studies and included neural tube defects, cardiovascular
defects, cleft defects, and chromosomal abnormalities. Exposure windows were selected to target
the period before or during gestation when exposure to a potential developmental toxicant or
mutagen might have the most profound effect on a particular developmental or genotoxic
endpoint. Maternal age, previous births to the mother, maternal smoking, and neighborhood
34
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family income were assessed as potential confounders. Exposure to BDCM at concentrations
>20 |ig/L was associated with increased risk of neural tube defects (adjusted relative risk 2.5;
95% confidence interval 1.2 to 5.1) when adjusted for maternal age and income level. However,
there was no evidence of a concentration-response trend. In addition, the study authors noted
that this point estimate was "fairly unstable" as a result of the low number of cases (n=10) in the
>20 |ig/L exposure category. There was no apparent trend or significant association for exposure
to BDCM and occurrence of cleft defects or chromosomal aberrations. A related earlier study
using similar exposure measurement methods and a cohort of 93,000 women who delivered
singleton births evaluated the relationship between pregnancy outcomes and TTHM exposure
(Dodds et al., 1999). This earlier study found an elevated relative risk for stillbirths in subjects
with exposure to >100 |ig/L TTHM during pregnancy (OR, but no significant increases in low
birth weight, very low birth weight, small for gestational age, pre-term delivery, or congenital
anomalies.
Shaw and colleagues investigated the association between tap water consumption and
cardiac defects (Shaw et al., 1990). They interviewed 145 mothers of children with severe
congenital cardiac disease and 176 mothers of children with no such disease. Subjects were asked
about their consumption of drinking water from tap and bottled sources during their first trimester
of pregnancy. A relationship was found between maternal consumption of 4 or more glasses of
tap water per day and increases in cardiac abnormality in the infants (OR 2.0, 95% CI 1.0-4.0). A
follow-up study was conducted by Shaw et al. (1991) to assess the relationship between water
chlorination and cardiac defects. In this study, the source of drinking water during the first
trimester was determined for each of the 138 study participants. THM concentrations
35
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corresponding to the first trimester of pregnancy were estimated from quarterly utility monitoring
records. No relationship between cardiac abnormalities and chlorinated drinking water was
identified, and average THM levels during the first trimester of pregnancy for cases (64.0 |ig/L)
was less than for controls (74.2 |ig/L).
Kramer et al. (1992) conducted a population-based case-control study, based on an
examination of birth certificates from January 1, 1989 to June 30, 1990 for live infants born to
white women in Iowa towns with 1,000 to 5,000 inhabitants. This study reported an increase in
intrauterine growth retardation with exposure to chloroform concentrations in drinking water of
greater than or equal to 10 |ig/L (OR 1.8, 95% CI 1.1-2.9). The authors also found a
nonsignificant increase in intrauterine growth retardation associated with exposure to drinking
water with BDCM concentrations greater than or equal to 10 |ig/L compared with drinking water
with undetectable BDCM concentrations (OR1.7, 95% CI 0.9-2.9). THM exposures, including
chloroform and BDCM, were estimated from a water supply survey conducted two to three years
previously.
King et al. (2000a) conducted a retrospective cohort study of 50,000 deliveries and
reported that exposure to TTHMs, chloroform and BDCM was associated with an increased
incidence of stillbirth. Exposure was estimated by linking the mother's residence at the delivery
date to the levels of THMs monitored in public water supplies, using the predicted average
exposure levels for the entire duration of the pregnancy. For chloroform, the adjusted odds ratio
for stillbirth was increased for exposure >100 |ig./L (OR 1.56, 95% CI 1.04-2.34). Risk doubled
for women exposed to a BDCM level of greater than or equal to 20 |ig/L, when compared to
women consuming concentrations of less than 5 |ig/L. When categories of stillbirth (unexplained
36
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deaths and asphyxia-related deaths) were examined, relative risk estimates for asphyxia-related
deaths increased by 32% for each 10 /i-g/L increase in concentration of BDCM. As indicated by
King et al. (2002a), an in vitro study (Alston 1991) suggested that chloroform and related
compounds might be linked to the asphyxia-related deaths by influencing the methionine-
homocysteine metabolic pathway and subsequent abruption of placenta. These data support the
asphyxia-related deaths observed by King et al. (2002a), and indicate that the observed stillbirths
were related to BDCM, despite their limitation to one type of stillbirth. A limitation of this study
is possible misclassification of exposure as a result of mobility within the study population.
Windham et al. (2003) examined menstrual cycle characteristics in relation to the presence
of brominated trihalomethanes in tap water in a prospective study of women living in Northern
California. Data were also reported for TTHMs and for chloroform. The target population was
married women of reproductive age (18-39 years old). Participants were selected from among
nearly 6500 women using a short screening interview to identify women who were more likely to
become pregnant (i.e., those who reported a menstrual period within the last six weeks, were not
surgically sterilized, did not use birth control pills or intra-uterine devices, and were non-
contracepting for less than 3 months). Out of 1092 eligible women, a total of 403 women finished
the study. These participants collected first morning urine samples daily for 2-9 menstrual cycles
(average 5.6 cycles) for measurement of steroid metabolites. The participants filled out a daily
diary during the urine collection phase and recorded vaginal bleeding. These measurements (diary
and urinary hormone metabolites) were used to estimate menstrual parameters such as cycle and
phase length. Cycle length was calculated from the first day of menses to the day before onset of
the next menses. When the available data permitted, the cycle was divided into the follicular
37
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phase (first day of menses through estimated day of ovulation) and the subsequent luteal phase.
Between 1424 and 1714 cycles were available for evaluation of each parameter. Information on
water consumption and other variables (age, race, education, employment, income, pregnancy
history, exercise type and frequency, smoking, alcohol and caffeine consumption) was collected in
a baseline telephone interview prior to urine collection. Information on the number of showers
taken at home per week and their duration was also collected. Exposure to trihalomethanes was
estimated from quarterly utility monitoring data and information on drinking water and other tap
water use collected during the baseline interview.
A monotonic decrease in mean cycle length was observed with increasing TTHM
exposure category. At TTHM concentrations greater than 60 |ig/L, the adjusted decrement in
cycle length was 1.1 day (95% C.I. -1.8, -0.40) when compared to TTHM concentrations of 40
|ig/L or less. The decrease in follicular phase length was similar (-0.94 day; 95% C.I. -1.6, -0.24).
A unit decrement in mean cycle and follicular phase length of 0.18 days per 10 |ig/L increase in
TTHM concentration (95% C.I. -0.29, -0.07) was determined when the cycle-specific TTHM
level was examined as a continuous variable. Examination of time spent showering did not reveal
additional risks with longer showers. Combined with TTHM concentration, decrements in cycle
and follicular phase length were seen at the higher TTHM (>60 |ig/L) and longer showers (>70
minutes) categories (-1.2 and -1.6 days respectively). However, the confidence intervals were
wide for all duration categories and a clear dose response pattern (i.e., shorter lengths at higher
durations) was not evident.
This study suggests that THM exposure may have effect on menstrual cycle length.
However, examination of time spent showering did not reveal additional risks with longer
38
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showers. This is counter to the expected trend, as elevated blood levels of THMs have been
documented after showering, due to dermal and inhalation exposure in the shower. However,
information on shower duration was collected by interview and the reported lengths may not have
accurately reflected actual shower duration. It would also be useful to have an independent
confirmation of these results in another study.
The weight-of-evidence from epidemiology studies suggests that DBPs are associated
with developmental/reproductive effects under certain exposure conditions. The existing data are
relatively sparse, and are insufficient for dose-response analysis. There are inconsistencies among
the available studies on the association between drinking water disinfection and specific effects,
such as changes in menstrual cycle length, fetal growth, fetal viability, and congenital
abnormalities. The studies employing the best exposure assessment found the strongest
association between fetal growth and fetal viability and chlorinated drinking water (Bove et al.,
2002). Evaluation of the studies is made more uncertain by the difficulties of exposure
assessment, including the different composition of D/DBPs at different locations, and the variation
in composition over time at a single location. The specific DBP(s) responsible for reproductive
and/or developmental effects is not known. The existing data suggest a relationship between
THMs, particularly BDCM, and developmental effects, but this may be related to a tendency
among investigators to report concentrations of THMs, and not the concentrations of other
DBPs. The available data suggest that additional studies on THMs and other DBPs are needed.
2.1.2. Systemic Effects
39
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There are no epidemiology studies addressing whether fetuses or children are more
sensitive than adults to systemic effects of chlorinated drinking water.
2.1.3. Carcinogenicity
The epidemiologic evidence regarding a relationship between drinking water chlorination
and cancer in adults is mixed. An association between chlorinated drinking water and rectal,
colon, kidney, and/or bladder cancers has been reported based on a number of human studies.
The association is strongest between DBPs and bladder cancer, but causality has not been
established. Koivusalo et al. (1998) and Cantor et al. (1998) both found evidence of increased
risk with increasing exposure duration, but the increase was only approximately two-fold after a
long exposure duration. An intermediate level of evidence supports an association with rectal
cancer. Yang et al. (1998) and Hildesheim et al. (1998) both found associations between
chlorinated drinking water exposure and rectal cancer, and the associations had a similar
magnitude in both sexes. Hildesheim et al. (1998) also found an association in both sexes with
lifetime average trihalomethanes (THMs) concentration. The consistency of the dose-response
trends, the consistency between sexes, and the apparent control of important potential
confounders in this study suggest that the observed associations between the exposures and rectal
cancer may be real (EPA, 2001e). Only one (King et al., 2000b) of three key recent studies
(Hildesheim et al., 1998; King et al., 2000b; Yang et al., 1998) found an association with colon
cancer, and a strong causal association between DBPs and colon cancer is considered unlikely
(EPA, 200 le). The limited data on kidney cancer support the possibility of an association with
DBPs. Yang et al. (1998) found fairly high standardized rate ratios for kidney cancer, and
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Koivusalo et al. (1998) observed a dose response by mutagenicity tertiles and by duration of
exposure. Limited human data (Cantor et al., 1999) also suggest an association between
chlorinated drinking water and gastrointestinal/urinary tract cancers and brain cancers (gliomas).
However, the data are insufficient for a definitive conclusion, and are insufficient for quantitative
cancer risk assessment.
Only one study specifically evaluated the relationship between childhood cancers and
chlorinated drinking water. In a population-based case-control study in Quebec, Infante-Rivard et
al. (2000) examined the possible association between childhood acute lymphoblastic leukemia
(ALL) and THMs. The authors studied 491 cases and 491 controls, matched for age, sex, and
region in the province; logistic regression analysis adjusted for maternal age and level of
schooling. Individual information was collected on water source, and exposure was estimated
from distribution system data for metals, nitrates, and THMs. No association with ALL was
found for prenatal or postnatal exposure to total THMs or for specific THMs, and there was no
evidence of a dose-response. Based on these studies, the data are inadequate to evaluate the
potential carcinogenic effects of childhood or prenatal exposure to chlorinated drinking water.
Overall, the epidemiology data are inadequate for a definitive conclusion regarding the
carcinogenic potential of chlorinated drinking water for adults or children.
2.2. TRIHALOMETHANES
2.2.1. Chloroform
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Chloroform is one of the best studied DBPs and has a very extensive toxicological
database. Chloroform and its metabolites have been shown to cause liver and kidney toxicity and
tumors as their primary adverse effects.
2.2.1.1. Developmental/Reproductive Effects
In a prospective study, Windham et al. (2003) demonstrated that increasing levels of
TTHMs were associated with significantly shorter cycles when examined by quartile. Similar
decrements were observed in follicular, but not luteal, phase length. However, the effect was
attributed to exposure to brominated THMs, and a clear association with reduced cycle length
was not observed for chloroform even at the highest quartile (>17 |ig/L) (difference -0.3 days;
95% C.I. -1.0, 0.40).
Two epidemiology studies (Kramer et al., 1992; King et al., 2000a) investigated the
relationship between exposure to chloroform in chlorinated drinking water and developmental
effects. These studies were discussed in Section 2.1., Chlorinated Drinking Water.
Three developmental toxicity studies (two in rats and one in rabbits) by the oral route of
administration (Thompson et al., 1974; Ruddick et al., 1983) and three developmental toxicity
studies (two in rats and one in mice) by the inhalation route of administration (Schwetz et al.,
1974; Murray et al., 1979) were reported.
Thompson et al. (1974) studied the effects of chloroform on embryonic and fetal
development of Sprague-Dawley rats. Groups of 25 pregnant rats (181-224 g) were gavaged
with chloroform in corn oil at total daily doses of 0, 20, 50, or 126 mg/kg/day by oral intubation
on days 6-15 of gestation, administered in two doses/day. Dams receiving 50 or 126 mg/kg/day
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displayed signs of maternal toxicity (decreased weight gain, mild fatty changes in the liver). There
was no evidence of maternal toxicity at 20 mg/kg/day, although microscopic examinations were
conducted on only 2 dams/group. Fetuses were removed by caesarean section 1 or 2 days prior
to expected parturition and examined for external, skeletal and/or soft tissue abnormalities. There
were no fetal malformations, but some fetal variations were noted as reflected in the statistically
significant increase in the incidence of bilateral extra lumbar ribs (p<0.05) at the high dose;
however, the increase in affected litters was not statistically significant. Fetal weight was also
reduced at the high dose (p<0.05). This study identified a maternal NOAEL of 20 mg/kg/day and
a LOAEL of 50 mg/kg/day in rats. For developmental effects, the NOAEL was 50 mg/kg/day,
with a LOAEL of 126 mg/kg/day.
In the same study, Thompson et al. (1974) administered chloroform (in corn oil) to
Dutch-Belted rabbits. In a preliminary range-finding study, doses of 0, 25, 63, 100, 159, 251, or
398 mg/kg/day were administered to pregnant rabbits on days 6-18 of gestation. High levels of
maternal death (60-100%) were observed at doses of 100 mg/kg/day and above. Adverse effects
at 63 mg/kg/day included anorexia, weight loss, diarrhea, abortion, and one maternal death. No
overt signs of toxicity other than mild diarrhea and intermittent anorexia were observed in dams
dosed with 25 mg/kg/day. In the main study, groups of 15 pregnant dams (1.7-2.2 kg) were
dosed by oral intubation with chloroform at 0, 20, 35, or 50 mg/kg/day on days 6-18 of gestation.
Decreased maternal weight gain was observed in dams given 50 mg/kg/day. Four high-dose dams
died from hepatotoxicity, but no evidence of hepatotoxicity was observed in surviving rabbits.
Four high-dose dams aborted, but this was not considered to be a treatment-related effect as three
control animals also aborted. Histopathology examinations revealed no evidence of maternal
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toxicity at 35 mg/kg/day. Small reductions in body weights (7.5% and 12%, respectively) were
observed in fetuses from dams administered 20 or 50 mg/kg/day (p<0.05), whereas only a 5.5%
decrease in fetal weight was observed at 35 mg/kg/day. At least some of the decrease in fetal
weight at the high dose may be attributable to the larger litter size at the high dose (7.4, vs. 6.4 in
the controls), although the mean litter size at the mid dose was only 4.5. An increased incidence
of fetuses with incompletely ossified skull bones (usually parietals) was observed at 20 and 35
mg/kg/day (p<0.05); a smaller increase at the high dose was not statistically significant, and the
results were not significant when the litter was used as the statistical unit of comparison. This
study is limited by the high incidence of abortions and mortality in the control group, but the study
authors did not consider the observed effects to be evidence of teratogenicity or fetotoxicity. This
study identified a maternal NOAEL of 35 mg/kg/day and a maternal LOAEL of 50 mg/kg/day,
based on hepatotoxicity; there were no developmental effects related to chloroform treatment.
Ruddick et al. (1983) investigated the developmental toxicity of chloroform in groups of
15 mated Sprague-Dawley rats. Pregnant dams (8-14 animals per dose group) were given 0, 100,
200, or 400 mg/kg chloroform in corn oil on days 6-15 of gestation. Maternal weight gain was
depressed by at least 20% at all dose levels. In addition, all dose levels of chloroform produced
maternal liver enlargement, decreased hemoglobin, and decreased hematocrit. Levels of serum
inorganic phosphorus and cholesterol were elevated in the dams at the highest exposure level.
Fetal weight was decreased by about 19% at the highest dose level. There were no fetal
malformations, but sternebra aberrations were observed with a dose-dependent incidence at 200
mg/kg/day and 400 mg/kg/day. Interparietal deviations also occurred at the high dose. There
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was a clear increase in the incidence of these variations indicating a potential developmental
effect.
In an inhalation study, Schwetz et al. (1974) exposed pregnant female Sprague-Dawley
rats to chloroform at target concentrations of 0, 30, 100 or 300 ppm (actual concentrations of 0,
30, 95, or 291 ppm; 0, 146, 464 or 1,420 mg/m3) for 7 hours/day from gestation days 6 through
15. Because marked anorexia was observed in an earlier experiment in dams exposed to 300 ppm
chloroform, an additional control group was starved; that is, allowed only 3.7 grams of food per
day. The numbers of pregnant rats exposed in each group were 68, 22, 23, and 3, respectively.
The low percent pregnancy observed at the high concentration was not considered to be
treatment-related, due to the time of exposure; however, the use of such a small number of
animals in the 300 ppm group decreased the statistical sensitivity of any adverse effects observed
in this group. The dams were sacrificed on gestation day 21, and fetuses were removed by
caesarian section. Food consumption and body weight gain exhibited concentration-related
decreases in all exposure groups. A significant increase in relative liver weights in dams exposed
to 100 or 300 ppm was observed at study termination, with a significant decrease in absolute liver
weight at 300 ppm. However, there was no effect on serum glutamate-pyruvate transaminase
activity at any concentration. At 300 ppm, 61% of the implantations were resorbed, a statistically
significant increase. This high resorption rate was not observed in the "starved" control group,
suggesting that weight loss cannot account for the observed effect, although the starved control
group was provided more food than was consumed by the 300 ppm group. Fetal body weights
were significantly decreased (40%) at 300 ppm, and fetal crown-rump lengths were slightly, but
significantly, decreased (2%) at 30 ppm and significantly decreased (15%) at 300 ppm. The
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frequencies of litters with acaudia or imperforate anus were significantly increased at 100 ppm.
Malformations were not observed at 300 ppm, but there were only three litters at this
concentration. The frequency of litters with delayed ossification was elevated in all exposure
groups. In addition, there were statistically significant increases in wavy ribs at 30 ppm, and in
missing ribs and subcutaneous edema at 100 ppm. The authors concluded that chloroform
exposures of 100 and 300 ppm were highly embryotoxic and fetotoxic, with embryolethality a
significant effect at 300 ppm.
Murray et al. (1979) found that 100 ppm chloroform was teratogenic in CF-1 mice
exposed on gestation days 8-15, but was fetotoxic in mice exposed on gestation days 1-7 or
6-15. Groups of 34-40 pregnant females (as determined by vaginal plug) were exposed to 0 or
100 ppm (0 or 490 mg/m3) for 7 hours/day on gestation days 1-7, 6-15, or 8-15, and sacrificed
on gestation day 18. The ability of the mice to maintain pregnancy was significantly decreased in
the groups exposed on gestation days 1-7 or 6-15, and there was a slight (but not statistically
significant) decrease in pregnancies in the group exposed on gestation days 8-15. Statistically
significant decreases in fetal weight and fetal length were observed in the groups exposed on
gestation days 1-7 and 8-15 but not on days 6-15. Cleft palate was observed at a statistically
significant increased incidence in litters of mice exposed on gestation days 8-15, but not in the
other groups. Cleft palate was seen predominantly in fetuses with retarded growth. No other
malformations were significantly increased in any group, although increased incidences of two
skeletal variations were observed. Delayed ossification of skull bones was significantly increased
in all exposed groups, and delayed ossification of sternebrae was significantly increased in the
groups exposed on gestation days 1-7 and 8-15, but not 6-15. The study authors suggested that
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the lack of malformations in the group exposed on gestation days 6-15 may have resulted from
the lethality to the early embryo obscuring other effects. Maternal toxicity was evident as
increased liver weight and increased serum glutamate-pyruvate transaminase activity.
Based on the findings of animal studies discussed above, developmental effects have been
found after chloroform exposure (e.g., pup weight reduction, skeletal variations). These prenatal
effects, however, were typically associated with exposures causing maternal toxicity, and
occurred at oral doses above those causing hepatotoxicity. The oral NOAEL for developmental
toxicity is in the range of 35-50 mg/kg/day, and the oral LOAEL for hepatotoxicity is 12.9
mg/kg/day for chloroform.
A multigeneration reproductive assay was conducted with chloroform in CD-I mice
(NTP, 1988). This assay evaluated reproductive effects in two successive generations, as well as
systemic effects in the second generation (i.e., Fx animals). In the first phase, mice were
administered chloroform by gavage in corn oil at 6.6, 16, or 41 mg/kg/day, 7 days/week for 18
weeks. In the second phase, the last litter of the control and of the high-dose groups were
retained. After weaning, the mice were administered the same chloroform dose as their parents,
and dosing continued through mating and parturition, when the study was terminated. No
adverse effects on fertility or reproduction of the Fx generation were observed, although increased
liver weight and liver lesions (degeneration of centrilobular hepatocytes, accompanied by
occasional single cell necrosis) were observed in all females exposed to the single dose tested.
The degeneration was characterized as minimal in 2/20, mild in 9/20 and moderate in 9/20
animals. Thus, a dose of 41 mg/kg/day caused mild to moderate liver histopathology in F,
females. No NOAEL can be identified for this effect, because the low- and mid-dose groups were
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not evaluated histopathologically. However, no adverse effects on fertility or reproduction were
found.
2.2.1.2. Systemic Effects
Numerous animal studies in several species (rats, mice and dogs) have shown that liver
and kidney toxicity are primarily target sites for the systemic effects of chloroform. Nasal toxicity
is also found in the rat following inhalation exposure. Organ toxicity (including liver and kidney
tumor response) following chloroform treatment vary with the exposure route, vehicle of
administration and strain of rat or mouse. The sensitivity to the organ toxicity induced by
chloroform is associated with oxidative metabolism. These results were summarized in two EPA
documents (EPA, 1994b, 1998b). Organ toxicity that results from chloroform is considered to be
part of the continuum that leads to tumor development. The organ toxicity is thus discussed
below in the context of the mode of carcinogenic action for chloroform.
2.2.1.3. Carcinogenicity
Chloroform has been found to cause liver and kidney tumors in rodents (discussed in EPA,
1994a, 1998b). A substantial body of data indicates that chloroform is not a DNA-reactive
mutagen. Thus, mutagenicity is not the key influence of chloroform on the carcinogenic process.
Chloroform induces liver and kidney tumors at doses that cause cell injury or organ toxicity.
Numerous studies have shown that organ toxicity and regenerative proliferation are associated
with tumorigenicity of chloroform, and thus are key steps in its carcinogenic mode of action
(EPA, 1998b; ILSI, 1997). To explore the issue of whether fetuses or children are at increased
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cancer risk compared with adults, the mode of carcinogenic action of chloroform in children
versus adults was examined. To address this question, the Agency evaluated the available data for
children and adults on chloroform metabolism and age-related differences in the rate of cell
proliferation (EPA, 1998b).
Organ toxicity from chloroform is dependent on oxidative metabolism primarily by
cytochrome P450 CYP2E1 (as discussed in EPA, 1998b). The oxidative metabolism of
chloroform generates highly tissue-reactive metabolites. One metabolite is phosgene, a highly
reactive dihalocarbonyl, and the other is the strong acid HC1. The very high reactivity of
phosgene prevents it from entering the nucleus and forming adducts with DNA. The chloroform
metabolites produce cytotoxicity (cell death) and regenerative hyperplasia (EPA, 1998b; ILSI,
1997). This process may lead to tumor development if sustained. Metabolism of chloroform via
a reductive pathway, if it occurs, could lead to the formation of free radicals and tissue damage,
but the reductive pathway is absent or minor under normal physiological conditions (EPA,
2001a).
Given that oxidative metabolism is key to the carcinogenic potential of chloroform, studies
on CYP2E1 in fetal and adult tissues were evaluated (EPA, 1998b). Studies in humans have not
shown consistent results; however, in those studies showing expression of CYP2E1, levels in
fetuses were lower than those in adults (Boutelet-Bochan et al., 1997; Carpenter et al., 1996;
Hakkola et al., 1998a; Vieira et al., 1996). Vieira et al. (1996) suggested that CYP2E1 activity
increases rapidly in the 24 hours after birth, and that activity level in children aged 1 to 10 years is
comparable to that of adults.
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Animal studies of CYP2E1 provide evidence of rapid induction of this gene soon after
birth (Song et al., 1986; Umeno et al., 1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990).
The study by Schenkman et al. (1989) indicated that CYP2E1 protein is present in low levels in
CD rat neonates, rises to a peak level at age 2 weeks and subsequently decreases to adult levels
by puberty. Analysis of protein levels quantified from western blots showed a maximum at 2
weeks with decreasing levels at 4 and 12 weeks. The protein level at 12 weeks was
approximately 50% of the level at 2 weeks. The authors did not provide a statistical analysis of
this result, but it appears from the error bars that the 2-week and 12-week levels (but not 4-week
levels) were significantly different. Song et al. (1986) conducted a similar analysis in Sprague-
Dawley rats and reported a rapid transcriptional induction of CYP2E1 (P450) within 1 week
following birth that remained elevated throughout 12 weeks. Enzyme activity followed a similar
pattern. Ueno and Gonzalez (1990) showed that extracts from 3-day-old and 12-week-old rat
liver, but not those from fetal or newborn rat liver, were able to generate significant CYP2E1
transcription in vitro. The ability of the extract to result in transcription of CYP2E1 was slightly
greater at 12 weeks.
Taken together, the animal studies do not provide conclusive evidence of an early period
of increased enzymatic activity. If, however, the twofold increase in CYP2E1 induction in
animals were seen in humans, its importance in terms of chloroform toxicity would depend on the
dose. Under low-dose conditions (e.g., much lower than the Km) it is possible that an increase in
the level of enzyme would not have any effect on active metabolite formation, because the amount
of chloroform, and not CYP2E1, would control the rate of phosgene production. On the other
hand, under saturating doses of chloroform, all the available enzyme would be active; thus a
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twofold increase in CYP2E1 could result in greater activation of the compound. Although the
animal data remain unclear regarding the potential for a neonatal period of increased CYP2E1
activity above that in the adult, the data in humans show a rapid induction after birth, gradually
increasing over the first year to reach adult levels during years 1-10 (Vieira et al., 1996).
Therefore, although children may have the capacity to metabolize chloroform, data on CYP2E1
activity provide no evidence to suggest that children have an increased susceptibility to
chloroform toxicity compared with adults (EPA, 1998b). Furthermore, the data from Schenkman
et al. (1989) indicate that levels of CYP2E1 are approximately 2-fold higher in rats at 2 weeks
than at 12 weeks of age. The 2-fold magnitude of the difference in CYP2E1 levels is within the
default UF value of 3 used to account for intraspecies variability in toxicokinetics (the CYP2E1
studies do not address the contribution of toxicodynamics to the total default UF of 10). Based
on this analysis, differences in age-dependent metabolism of chloroform by CYP2E1 are
adequately accounted for in the existing UF for intraspecies variability.
Thus, the human data are inconclusive on age-related differences in CYP2E1 activity, and
it is not possible to state whether children may have an increased susceptibility to chloroform
toxicity and carcinogenicity as compared to adults. However, based on animal data, the degree of
age-dependent differences in CYP2E1 expression (if any) are likely well accounted for by default
assumptions about the magnitude of intraspecies variability.
The next issue to examine is whether the developing fetus may be more sensitive to the
toxicity of chloroform because of its greater rate of cell proliferation. There are very few data on
prenatal and postnatal exposures to chloroform and resultant organ toxicity. Liver toxicity was
found in a multigeneration reproductive assay in CD-I mice (NTP, 1988). The periods of
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exposure included prenatal, postnatal and adult stages. The liver toxicity from this
multigeneration reproductive study was compared with that of a comparable 90-day study in adult
B6C3F1 mice for liver toxicity (Bull et al., 1986). The similarity of effects at comparable doses
from these two studies suggests that there is no increased susceptibility to chloroform that results
from prenatal or postnatal exposures (EPA, 1998b). It should be noted that there are limitations
in this comparison; different strains of mice were used, and only LOAELs were identified in these
two studies (EPA, 1998b).
Under the 1986 U.S. EPA Guidelines for Carcinogen Risk Assessment, chloroform has
been classified as Group B2, probable human carcinogen, based on sufficient evidence of
carcinogenicity in animals (EPA, 1998a). Under the Draft Guidelines for Carcinogen Risk
Assessment (EPA 1999), chloroform is likely to be carcinogenic to humans by all routes of
exposure under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in
susceptible tissues (EPA, 2001a). Chloroform is not likely to be carcinogenic to humans by any
route of exposure under exposure conditions that do not cause cytotoxicity and cell regeneration.
This weight-of-evidence conclusion is based on several lines of evidence. 1) Observations in
animals exposed by both oral and inhalation pathways indicate that sustained or repeated
cytotoxicity with secondary regenerative hyperplasia precedes, and is probably required for,
hepatic and renal neoplasia. 2) There are no epidemiological data specific to chloroform. The
epidemiological data relating cancer to exposure to chlorinated drinking water are mixed, and any
observed effects cannot necessarily be attributed to chloroform amongst multiple other
disinfection byproducts. 3) Genotoxicity data on chloroform are essentially negative; there are a
few scattered positive results that generally have limitations such as excessively high doses or
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confounding factors. Thus, the weight-of-evidence of the genotoxicity data on chloroform
supports a conclusion that chloroform is not strongly mutagenic, and that genotoxicity is not
likely to be the predominant mode of action underlying the carcinogenic potential of chloroform
(EPA, 2001a).
Because a substantial database indicates that tumor development for chloroform is
secondary to organ toxicity and regenerative proliferation, a nonlinear dose-response approach for
tumorigenicity is viewed as appropriate. Use of a low dose-linear approach is considered overly
conservative for extrapolating cancer risk associated with chloroform exposure (EPA, 1998a,b,
2001a; ILSI, 1997). EPA (2001a) determined a point of departure (POD) of 23 mg/kg/day,
based on increased kidney tumors in Osborne-Mendel rats (Jorgensen et al., 1985). The Agency
compared this POD to the RfD of 0.01 mg/kg/day calculated as described in the next paragraph,
and concluded that the margin of exposure (MOE) of 2000 was adequately protective of public
health for cancer effects. Sustained tissue toxicity, which is a key event in the cancer mode of
action for chloroform, will not occur at doses below the RfD.
2.2.1.4. Basis for RfD and MCLG
The Agency used an oral study in dogs (Heywood et al., 1979) to derive the chloroform
RfD. The study identified a LOAEL of 15 mg/kg/day for hepatotoxicity (fatty cysts and an
increase of serum glutamic pyruvic transaminase). A BMDL10 (the 95% lower bound confidence
limit on the dose associated with a 10% extra risk) of 1.2 mg/kg/day was also calculated, based
on the prevalence of animals with moderate to marked fatty cysts in liver in the same study.
Applying a factor of 6/7 to account for exposure 6 days/week, the LOAEL was converted to 12.9
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mg/kg/day, and the BMDL10 was converted to 1.0 mg/kg/day. The RfD was based on both the
benchmark dose (BMD) and traditional NOAEL/LOAEL approach, which coincidentally result in
the same RfD. An RfD of 0.01 mg/kg/day was calculated using the BMD approach by applying
an overall UF of 100 (10 for interspecies extrapolation and 10 for protection of sensitive
individuals) to the BMDL10 of 1.0 mg/kg/day. Using the traditional approach, a composite UF of
1,000 was applied (100 was used to account for inter- and intraspecies differences and a factor of
10 for use of a LOAEL), resulting in an RfD of 0.01 mg/kg/day. This RfD corresponds to a
Drinking Water Equivalent Level (DWEL) of 0.35 mg/L, assuming an adult tap water
consumption of 2 L/day for a 70 kg adult.
Because a nonlinear approach was used for chloroform carcinogenicity, the MCLG is
based on liver toxicity as the most sensitive effect for chloroform (as it is the lowest possible
LOAEL for any organ) and as a precursor response to a key step to its carcinogenicity. This
approach is considered equally protective of both adults and children because the database on
chloroform does not indicate that children are more sensitive than adults to liver toxicity. The
mode of action by which chloroform produces organ toxicity and carcinogenicity is considered to
be the same for children and adults.
The MCLG of 0.07 mg/L was calculated from the RfD by assuming an adult tap water
consumption of 2 L per day for a 70 kg adult, and by applying a relative source contribution
(RSC) of 20%. The RSC is based on data indicating that exposure to chloroform by other routes
and sources of exposure may potentially contribute a substantial percentage of the overall
exposure to chloroform (EPA, 200Id). Based on average daily doses for each source and route
of exposure under specific conditions, EPA estimated that for the median individual, ingestion of
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total tap water (including water-based drinks) can contribute roughly 28% of the total dose of
chloroform, while the dose from showering (inhalation and dermal exposure) was estimated to
contribute approximately 14% of the total dose (EPA, 2001d). There is, however, considerable
uncertainty in these exposure estimates. For example, because chloroform is so volatile, most of
the chloroform would evaporate from ingested hot liquids such as coffee or tea, so the value of
28% from ingestion of total tap water may be an over-estimate. In addition, the proportion of
intake from different sources varies widely across the population. Therefore, the default RSC of
20% is used, a value that is consistent with the percent of the total dose from ingestion of total
tap water. Thus, the MCLG is 0.07 mg/L: (0.01 mg/kg/day x 70 kg x 0.2)/ (2/L/day) = 0.07
mg/L.
2.2.1.5. Children's Risk in Relation to the MCLG
The MCLG derived for chloroform is considered protective of both adults and children,
given that developmental effects occurred at doses above those causing hepatotoxicity. In
addition, fetal effects were only seen at levels at which maternal toxicity was noted. Also, the
mode of action data indicates that children are not uniquely sensitive to the organ toxicity caused
by high doses of chloroform and there is no evidence from the available studies to suggest that
children or fetuses would be qualitatively more sensitive to its effects than adults. In addition, the
developing fetus would not be expected to be particularly sensitive to a cytotoxic agent, such as
chloroform at low levels, because cell division occurs at a rapid pace and the cellular repair
capacity is high (EPA, 2001a).
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The MCLG is also considered protective of carcinogenic effects in children and fetuses.
EPA believes that using a nonlinear dose-response approach for setting the MCLG is protective
for children and fetuses for the following reasons: 1) the reactive metabolite inside the cell should
have similar effects resulting from its reacting with and disrupting macromolecules in the cells of
fetuses, children, and adults; 2) cell necrosis and reparative replication are not likely to be
qualitatively different in various life stages; 3) cancer risk to the fetus or children would be a
function of cytotoxicity, as in adults, and protecting fetuses and children from sufficient levels of
the chemical that would cause cytotoxicity should protect against cancer risk (EPA, 2001a); 4) a
comparison of liver toxicity between a multigeneration reproductive study and a comparable 90-
day study suggests that there is no increased susceptibility to chloroform effects on the liver that
results from prenatal or postnatal exposure, indicating that children are not more sensitive to the
liver cytotoxicity that is a precursor to chloroform carcinogenicity (EPA, 1998b). This final
factor also supports the conclusion that children are not more sensitive than adults to the
noncancer liver effects of chloroform.
2.2.2. Brominated Trihalomethanes
There is sufficient evidence for carcinogenicity via ingestion of bromoform and
bromodichloromethane (BDCM) to consider them probable human carcinogens. The evidence is
limited for dibromochloromethane (DBCM). Based on the available data, a mechanism of action
involving mutagenicity was postulated for the brominated THMs, indicating that linear low-dose
extrapolation for BDCM and bromoform is appropriate. The proposed mechanism of
carcinogenicity for these compounds was examined to determine if this would provide any reason
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for concern that children or fetuses may be more susceptible to development of cancer following
exposure. If carcinogenicity is the result of mutations by either the parent compound or a
metabolite, children or the developing fetus could be more susceptible to the carcinogenicity of
brominated THMs due to a higher rate of cell proliferation in the target organs. An increased risk
of this type would be true for all genotoxic carcinogens and not specific to brominated THMs.
There are no data currently available for brominated THMs to permit quantification of a possible
increase in risk to the developing fetus or children.
A number of epidemiology studies have reported an association between exposure to
THMs and developmental/reproductive effects (Bove et al., 1995; Dodds et al., 1999; Gallagher
et al., 1998; King et al., 2000a; Klotz and Pyrch, 1999; Kramer et al., 1992; Savitz et al., 1995;
Waller et al., 1998, Windham et al., 2003). For all of these studies, because the subjects were
exposed to other contaminants and disinfection byproducts in the drinking water, correlation of
the effects directly to individual brominated THM exposure is difficult. Further details about
these studies were presented in Section 2.1, Chlorinated Drinking Water.
Although the mechanism of brominated trihalomethane toxicity is not known with
certainty, data indicate that the adverse effects of this group of chemicals are secondary to
metabolism. Brominated THMs are extensively metabolized via oxidative (using NADPH and
oxygen) and reductive (using NADPH or NADH and inhibited by oxygen) pathways in humans
and animals, primarily in the liver, but also in the kidney (EPA, 2002c). Recent data suggest that
bioactivation of brominated trihalomethanes to mutagenic species is also mediated by one or more
glutathione .S'-transferase-mediated conjugation pathways.
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Both oxidative and reductive reactions are mediated by cytochrome-P450s. Oxidative
metabolism results in the production of a dihalocarbonyl (CX20) intermediate, which may
undergo a variety of reactions, such as adduct formation with cellular nucleophiles, hydrolysis to
yield carbon dioxide or glutathione-dependent reduction to yield carbon monoxide. Reductive
metabolism results in the production of free radical species such as the dihalomethyl radical
(CHX2 ). In vitro and in vivo data suggest that metabolism via the reductive pathway occurs
more readily for the brominated trihalomethanes than it does for chloroform. Gao and Pegram
(1992) reported that binding of reactive intermediates to rat hepatic microsomal lipid and protein
under reductive (anaerobic) conditions was more than twice as high for bromodichloromethane as
for chloroform. Tomasi et al. (1985) used electron spin resonance (ESR) spectroscopy to
measure the production of a free radical intermediates (a product of the reductive pathway) in
vitro using rat hepatocytes isolated from phenobarbital-induced male Wistar rats. The intensity of
the ESR signal was greatest for bromoform, followed by BDCM and then chloroform. The
largest ESR signal was detected when hepatocytes were incubated under anaerobic conditions.
Tomasi et al. (1985) also used ESR to evaluate free radical production in vivo in rats given
intraperitoneal injections of chloroform, BDCM, or bromoform. The intensity of the ESR signal
followed a ranking similar to that observed in in vitro experiments (bromoform >BDCM>
chloroform), confirming that the reductive formation of free radicals is greater for brominated
trihalomethanes than for chloroform. Together, these data indicate that reductive metabolism is a
more important pathway for metabolism of brominated trihalomethanes than for chloroform. The
relative importance of the oxidative and reductive pathways for the brominated THMs in vivo has
not been determined, but it is of note that oxygen partial pressure in the kidney and liver is low
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under in vivo conditions. Both dihalocarbonyls (produced by the oxidative pathway) and
dihalomethyl radicals (produced by the reductive pathway) are reactive species and may form
covalent adducts with a variety of cellular components (EPA, 2002c).
Metabolism of the brominated THMs involves a number of enzymes, including CYP2E1,
CYP2B1/2 and CYP1A (EPA, 2002c). The roles of the different P450 isozymes in brominated
trihalomethane metabolism (i.e., oxidative versus reductive pathways) have not been definitively
identified, but CYP2E1, CYP2B1/2 and CYP1A have been implicated in the oxidative pathway.
The toxicity of BDCM and DBCM has been shown to be at least partly related to bioactivation by
the enzyme CYP2E1 (e.g., Thornton-Manning et al., 1994). Thus, a higher level of CYP2E1
activity in children, as compared to adults, would mean that children could be at greater risk for
carcinogenic effects from these compounds.
Studies in humans have not shown consistent results; however, in those studies showing
expression of CYP2E1, levels in fetuses were lower than those in adults (Boutelet-Bochan et al.,
1997; Carpenter et al., 1996; Hakkola et al., 1998a; Vieira et al., 1996). Vieira et al. (1996)
suggested that CYP2E1 activity increases rapidly in the 24 hours after birth, and that activity level
in children aged 1 to 10 years is comparable to that of adults.
Animal studies of CYP2E1 provide evidence of rapid induction of this gene soon after
birth (Song et al., 1986; Umeno et al., 1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990).
The study by Schenkman et al. (1989) indicated that CYP2E1 protein is present in low levels in
CD rat neonates, rises to a peak level at age 2 weeks and subsequently decreases to adult levels
by puberty. Analysis of protein levels quantified from western blots showed a maximum at 2
weeks with decreasing levels at 4 and 12 weeks. The protein level at 12 weeks was
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approximately 50% of the level at 2 weeks. The authors did not provide a statistical analysis of
this result, but it appears from the error bars that the 2-week and 12-week levels (but not 4-week
levels) were significantly different. Song et al. (1986) conducted a similar analysis
in Sprague-Dawley rats and reported a rapid transcriptional induction of CYP2E1 (P450) within
one week following birth that remained elevated throughout 12 weeks. Enzyme activity followed
a similar pattern. Ueno and Gonzalez (1990) showed that extracts from 3-day-old and 12-week-
old rat liver, but not those from fetal or newborn rat liver, were able to generate significant
CYP2E1 transcription in vitro. The ability of the extract to result in transcription of CYP2E1
was slightly greater at 12 weeks.
Taken together, the animal studies do not provide conclusive evidence of an early period
of increased enzymatic activity. If, however, the 2-fold increase in CYP2E1 induction in animals
were seen in humans, its importance in terms of brominated THM toxicity would depend on the
dose. Under low-dose conditions (e.g., much lower than the Km) it is possible that an increase in
the level of enzyme would not have any effect on active metabolite formation, because the amount
of brominated THMs, and not CYP2E1, would control the rate of the enzymatic metabolism. On
the other hand, under saturating doses of brominated THMs, all the available enzyme would be
active; thus a 2-fold increase in CYP2E1 could result in greater activation of the compound.
Although the animal data remain unclear regarding the potential for a neonatal period of increased
CYP2E1 activity above that in the adult, the data in humans show a rapid induction after birth,
gradually increasing over the first year to reach adult levels during years 1-10 (Vieira et al., 1996).
Therefore, although children may have the capacity to metabolize brominated THMs, data on
CYP2E1 activity provide no evidence to suggest that children have an increased susceptibility to
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brominated THM toxicity compared with adults (EPA, 2002c). Furthermore, the data from
Schenkman et al. (1989) indicate that levels of CYP2E1 are approximately 2-fold higher in rats at
2 weeks versus 12 weeks of age. The 2-fold magnitude of the difference in CYP2E1 levels is
within the default UF value of 3 used to account for intraspecies variability in toxicokinetics (the
CYP2E1 studies do not address the contribution of toxicodynamics to the total default UF of 10).
Based on this analysis, differences in age-dependent metabolism of brominated THMs by CYP2E1
are adequately accounted for in the existing UF for intraspecies variability.
Thus, the human data are inconclusive on age-related differences in CYP2E1 activity and
it is not possible to state whether children may have an increased susceptibility to brominated
THM toxicity as compared to adults. However, based on animal data, the degree of age-
dependent differences in CYP2E1 expression (if any) are likely well accounted for by default
assumptions about the magnitude of intraspecies variability.
Information on age-related differences in the activity of the other human cytochrome
P450s involved in brominated THM metabolism was reviewed in a recent summary of the current
literature (EPA, 2000d). For CYP1 Al, constitutive fetal mRNA levels were measurable in some
studies, with expression levels continuing to increase through gestation (Oesterheld, 1998).
However, a decline in CYP1 Al activity was reported following birth (Rendic and Di Carlo,
1997). No data on age-dependent differences in inducible levels of CYP1 Al were presented in
the review. For CYP1A2, low fetal mRNA levels and low enzyme activity in the neonate were
reported. However, there were similar levels of CYP1A2 enzyme activity in adults and in infants,
and activity was greater in young children than in adults (Hakkola et al., 1998a; 1998b). Based
on patterns of theophylline clearance as a measure of CYP1A2 activity, CYP1A2 is minimally
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active in fetuses, reaches maximum activity during early childhood and decreases thereafter (EPA,
2000d).
The human CYP isoforms CYP2A6, CYP2D6, and CYP3A4 are potential candidates for
metabolism of brominated THMs, based on overlapping catalytic activity with rodent CYP2B1/2
(WHO, 2000), but the identity of the human CYP isoforms, other than CYP2E1, capable of
metabolizing brominated THMs is not known. In light of this knowledge gap, the toxicological
consequences of their developmental expression patterns are not clear. However, except for
CYP1A2, fetuses and children generally have lower expression than adults of the CYPs that may
be involved in brominated THM metabolism (Hakkola et al., 1998a, 1998b; Tanaka, 1998;
Oesterheld, 1998).
Higher enzyme activity will not necessarily result in higher tissue doses of metabolite,
since at low doses (i.e., much lower than the Km), the amount of metabolism would be controlled
by the amount of substrate, not the amount of enzyme. In addition, if metabolism of the
brominated THMs by CYP1A, CYP2B1/2, or their analogues represents a low affinity route of
metabolism (as it does for chloroform), the impact of age-dependent expression of these enzymes
at environmental exposure levels may be less important than for CYP2E1. No data were
identified on the relative affinity of the brominated THMs for the different CYP isoforms.
Overall, the P-450 enzyme expression and activity data do not suggest that children would be
more susceptible than adults to brominated THM.
Only minimal data are available regarding the enzymes involved in the reductive
metabolism of brominated trihalomethanes. While experimental evidence indicates that CYP2E1
and CYPB1/2 catalyze the oxidative pathway, the identities of the cytochrome P450 isoforms that
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catalyze the reductive pathway have not been established. In general, CYP2E1 protein can
catalyze reductive as well as oxidative reactions (Lieber, 1997) and this isoform has been
implicated in the production of trichloromethyl radicals from carbon tetrachloride (see Lieber et
al. 1997). However, evidence for a dual role of either CYP2E1 or CYP2B1/2 in catalyzing the
oxidative and reductive pathways for trihalomethane metabolism has been contradictory, perhaps
as a result of the different concentrations of chloroform used in different experiments
(summarized in Testai et al., 1996). To address the issue of concentration, Testai et al. (1996)
studied the role of different isoforms in chloroform metabolism, and they found that the
cytochrome P450 isoforms involved in oxidative metabolism of brominated trihalomethanes do
not participate in the reductive pathway. Thus, no conclusion can be made on whether children
are more sensitive than adults based on the limited data on the reductive pathway.
The information on age-related differences in the production of metabolites via the
glutathione conjugation pathway is also very limited. Studies in Salmonella typhimurium strains
engineered to express the rat glutathione ^-transferase theta 1-1 (GSTT1-1) gene indicate that
metabolism of brominated THMs to mutagens may also be catalyzed by glutathione S-transferase
(DeMarini et al., 1997; Landi et al., 1999). Data on the GST theta genes are currently quite
limited; however, one study reported that theta-class GSTs were expressed in human adult liver,
but not fetal liver (Mera et al., 1994). These results suggest that the fetus does not experience
increased risk from GST theta-mediated mutagenicity. The occurrence of increased risk in
children cannot be evaluated, since the age at which expression of GST theta begins is unknown.
2.2.2.1. Bromodichloromethane
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Developmental/Reproductive Effects
Developmental and reproductive toxicity data are available for bromodichloromethane
(BDCM) and were considered in the derivation of the MCLG. Five epidemiologic studies have
reported a relationship specifically between developmental/reproductive toxicity and BDCM.
These studies were summarized in detail in Section 2.1.1., but key conclusions are highlighted
here. In a prospective study, Windham et al. (2003) demonstrated that increasing levels of
individual brominated trihalomethanes or total brominated trihalomethanes in the drinking water
were associated with significantly shorter cycles when examined by quartile. Similar decrements
were observed in follicular, but not luteal, phase length. For BDCM, the adjusted decrements
were 0.74 days (95% C.I. -1.5, -0.02) for mean cycle length and 0.8 days (95% C.I. -1.5, -0.08)
for mean follicular phase length at the highest quartile (>16 //g/L). The strongest association for
an individual compound was observed for DBCM. Menses length was slightly increased at the
highest quartile for BDCM exposure.
A population based case-control study (Kramer et al., 1992) reported an increased risk
(not statistically significant) of intrauterine growth retardation with exposure to BDCM
concentrations in drinking water of greater than or equal to 10 |ig/L, compared with drinking
water with undetectable BDCM concentrations (OR = 1.7, 95% CI = 0.9-2.9). As previously
discussed, Waller et al. (1998) reported that pregnant women exposed to TTHMs in drinking
water at levels of 75 //g/L or higher had an increased risk of spontaneous abortion. This study
also reported that consumption of five or more glasses of cold water with a BDCM concentration
of at least 18 |ig/L was associated with an increased risk of spontaneous abortion. After
adjustment for exposure to other THMs, the adjusted OR was 3.0 (CI = 1.4-6.6).
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King et al. (2000a) reported that exposures to TTHMs, chloroform and BDCM were
associated with an increased risk of stillbirth. Risk doubled for women exposed to a BDCM level
of greater than or equal to 20 //g/L, when compared to women consuming concentrations of less
than 5 //g/L. In a retrospective cohort study among 49,842 residents of Nova Scotia (Dodds and
King, 2001), exposure to BDCM at concentrations >20 ixgfL was associated with an increased
risk of neural tube defects (adjusted relative risk = 2.5, 95% CI = 1.2 - 5.1), but there was no
evidence of a dose-response trend.
Seven oral studies in laboratory animals evaluated the developmental toxicity of BDCM
(Bielmeier et al., 2001; CCC, 2000a, b, c; Narotsky et al., 1997a; NTP, 1998a; Ruddick et al.,
1983). To determine the effect of vehicle on BDCM toxicity, Narotsky et al. (1997a)
administered BDCM by gavage in either corn oil or an aqueous vehicle with Emulphor® to
pregnant F344 rats (12-14/group) at dose levels of 0, 25, 50, or 75 mg/kg/day during gestation
days 6-15. Decreased maternal weight gain and full-litter resorption were observed at 50 and 75
mg/kg/day. The incidence of full-litter resorption was significantly higher in the corn oil vehicle
(83%) compared with the aqueous vehicle (21%) at the high dose. Accordingly, the NOAEL for
developmental toxicity (full-litter resorptions) was 25 mg/kg/day, with a LOAEL of 50
mg/kg/day. The LOAELs for maternal effects (based on decreased maternal weight gain during
GD 6-8) in aqueous and corn oil vehicles were 25 and 50 mg/kg/day, respectively. No maternal
NOAEL for the aqueous study could be identified.
Bielmeier et al. (2001) conducted a series of experiments on the effect of BDCM on
pregnancy loss, characterized as full litter resorptions, in female F344 and Sprague-Dawley rats.
In one experiment, doses of 0, 75, or 100 mg/kg/day were administered via gavage to pregnant
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rats on gestation day 9 in 10% Emulphor. The dose-related incidence of full litter resorptions was
0% (0/8), 64% (7/11), and 90% (9/10), respectively, identifying a LOAEL of 75 mg/kg/day. This
effect was seen in F344 rats, but not in Sprague-Dawley rats. Dosing during different portions of
pregnancy showed that the critical period for induction of full litter resorption was limited to the
luteinizing hormone (LH)-dependent phase of pregnancy, suggesting that BDCM may disrupt
pregnancy via a LH-mediated mode of action. Measurement of serum LH and progesterone
levels indicated that full litter resorption was accompanied by a marked reduction in progesterone
concentration without a corresponding drop in LH levels. The failure of BDCM to exert adverse
effects after the LH-dependent window, the reduction in serum progesterone level, and the
unchanged serum LH levels led this group to conclude that the target of toxicity was the ovary
and that the mode of action was a reduced sensitivity of the corpus luteum to LH. Although there
are significant differences between rats and humans in the hormonal maintenance of pregnancy,
the authors did consider their findings possibly relevant to humans. Additional research, some of
which is currently ongoing, is needed to evaluate the relevance to humans of these findings in rats.
Ruddick et al. (1983) investigated developmental toxicity in pregnant Sprague-Dawley
rats (9-15/group) administered BDCM by gavage in corn oil at dose levels of 0, 50, 100, or 200
mg/kg/day from gestation days 6-15. Maternal weight gain was significantly depressed in the
high-dose group, and reduced in the low- and mid-dose groups; the decreases were not
statistically significant. There were no fetal malformations, but a dose-dependent increased
incidence of sternebra aberrations was observed in all dose groups. Although there was a clear
increase in the incidence of these variations, no statistical analysis was performed by the authors.
However, a statistical analysis (using the Fisher Exact Test) conducted on the published data
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found that none of these increases differed significantly from controls. A trend test showed a
statistically significant dose-related trend (p=0.03), and stepwise analysis indicated that the trend
became significant only when the high-dose (200 mg/kg/day) was included in the analysis. These
findings suggest that the NOAEL and LOAEL were 100 mg/kg/day and 200 mg/kg/day,
respectively. However, due to the small sample sizes, the statistical power of the experiment to
detect effects at lower doses is limited. The NOAEL and LOAEL for maternal effects were 100
mg/kg/day and 200 mg/kg/day, respectively, based on significantly decreased maternal weight
gain (EPA 2002c).
NTP (1998a) conducted a short-term study screening study in Sprague-Dawley rats
investigating both reproductive and developmental toxicity from BDCM administered in drinking
water. Two groups of male rats and three groups of female rats were treated with BDCM at
concentrations of 0, 100, 700, or 1300 mg/L. Based on measured water consumption, the
authors estimated dose levels for the treated males to be 0, 8, 41, or 68 mg/kg/day, and for the
treated females to be 0, 14, 72, or 116 mg/kg/day (groups A and C) and 0, 13, 54, or 90
mg/kg/day (group B). The rats were exposed for 25 to 30 days, with the exception of group B
females, which were exposed from gestation day 6 to evidence of littering/birth (approximately 15
days). BDCM exposure did not affect any reproductive parameter investigated in males or
females, with the exception of a non-dose related increase in the number of live fetuses per birth
at the 14 mg/kg/day dose in Group C females and a slight decrease in the number of live fetuses
per birth at the 72 mg/kg/day dose in Group A females. On the basis of these results, NTP
(1998a) concluded that BDCM was not a short-term reproductive or developmental toxicant at
doses up to approximately 68 and 116 mg/kg/day in male and female rats, respectively. The study
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sensitivity was decreased by the use of a relatively small number of animals per group (5-
13/sex/dose) and the lack of microscopic examination of the pups.
The Chlorine Chemistry Council sponsored a range-finding reproductive/developmental
toxicity study of BDCM in Sprague-Dawley rats (CCC, 2000a; Christian et al., 2001b). BDCM
was administered to parental rats (P generation, 10/sex/group) in drinking water at concentrations
of 0, 50, 150, 450, or 1350 ppm. Exposure began 14 days before cohabitation and continued
until the day of sacrifice. Lactation was extended for one week (LD 22-29) beyond the normal 3-
week period because F, pup body weights in the three highest dose groups were significantly
reduced on LD 21 relative to control values (results are described below). On LD 29, two F,
pups per sex were selected from each litter for an additional week of postweaning observation,
provided ad libitum access to water containing the same concentration of BDCM administered to
their parents (P generation), and sacrificed on Day 8 postweaning.
Exposure-dependent reductions in both absolute (g/day) and relative (g/kg body weight-
day) water consumption were observed in all rats of both sexes and were attributed to taste
aversion. Treatment-related clinical signs (e.g., dehydration, emaciation, chromorhinorrhea) were
observed in both sexes in the 1350 ppm exposure groups and were considered to be generally
associated with reduced water consumption. The only other observed effect was a concentration-
dependent reduction in F, pup body weights and weight gain in the 150, 450, and 1350 ppm
exposure groups, both prior to, and after weaning. Based on decreased pup weight and pup
weight gain, the LOAEL for developmental toxicity is 150 ppm, and the corresponding NOAEL
is 50 ppm. Although the effect of reduced water consumption on the decreases in feed
consumption, body weight gain, and body weight observed in the P generation adults is unclear,
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the LOAEL for parental toxicity is considered to be 150 ppm and the NOAEL is 50 ppm. Due to
the marked changes in drinking water consumption by P generation female rats during different
physiological stages (pre-mating, mating, gestation, and lactation), it is not possible to convert the
administered drinking water concentrations into a single biologically meaningful average daily
dose; instead, dose in mg/kg/day was reported separately for the premating, gestation, lactation,
and post-weaning phases of the study.
In the full study, the Chlorine Chemistry Council (CCC, 2000b; summarized in Christian et
al., 2001a) examined the developmental effects of BDCM in Sprague-Dawley rats. Pregnant rats
were given BDCM in drinking water at concentrations of 0, 50, 150, 450, or 900 mg/L (0, 2.2,
18.4, 45.0, or 82.0 mg/kg/day) on gestation days 6-21. No treatment-related clinical signs or
necropsy results were observed. Significantly reduced water consumption was observed in all
dose groups and was attributed to taste aversion. Decreases in absolute and relative feed
consumption were observed in the three highest dose groups; associated decreases in maternal
body weight gain were attributed to taste aversion. The effect on maternal body weight gain was
persistent at the two highest doses, but was transient at the lower doses. No effects on early or
late resorptions, fetal body weight, liver litter size, or other developmental parameters were
observed from BDCM exposure. There were no instances of full litter resorption and no dead
fetuses. The only statistically significant changes in the occurrence of skeletal variations were
reversible delays in ossification at the high dose, including an increased fetal incidence of wavy
ribs and a decreased number of ossification sites per fetus per litter for the forelimb phalanges and
the hindlimb metatarsals and phalanges. The study authors did not consider the increased fetal
incidence of wavy ribs to be related to BDCM exposure, because the more relevant measure of
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litter incidence did not differ significantly from the controls and was within the historical range for
this alteration at the test facility. The maternal NOAEL and LOAEL were 18.4 mg/kg/day and
45.0 mg/kg/day, respectively, based on significant reductions in maternal body weight and body
weight gains, while the developmental NOAEL and LOAEL were 45.0 mg/kg/day and 82.0
mg/kg/day, respectively, based on a significant decrease in the number of ossification sites per
fetus for the forelimb phalanges and the hindlimb metatarsals and phalanges.
The Chlorine Chemistry Council also examined the developmental effects of BDCM in
New Zealand white rabbits (CCC, 2000c). Pregnant rabbits were given BDCM in drinking water
at concentrations of 0, 15, 150, 450, or 900 mg/L (0, 1.4, 13.4, 35.6, or 55.3 mg/kg/day) on
gestation days 6-29. No treatment-related clinical signs or necropsy results were observed.
Significantly reduced feed and water consumption and body weight gains were observed in the
35.6 and 55.3 mg/kg/day dose groups. No observable effects on fetal body weight, live litter size,
and a number of other developmental parameters were noted from BDCM exposure. Statistically
significant increases in the number of fused sterna centra were observed in the 13.4 and 35.6
mg/kg/day dose groups; however, this effect was not dose-related, and the observed incidences
were within the historical range for the testing facility. The maternal NOAEL and LOAEL
identified in this study were 13.4 and 35.6 mg/kg/day, respectively, based on decreased body
weight gain, while the developmental NOAEL was 55.3 mg/kg/day, based on an absence of
statistically significant, dose-related effects at any dose tested.
Christian et al. (2002) conduced a standard two-generation reproductive toxicity study in
which BDCM was continuously provided to Sprague-Dawley rats in the drinking water at
concentrations of 0, 50, 150, or 450 ppm. Average daily doses estimated by the study authors
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were 4.1 to 12.6, 11.6 to 40.2, and 29.5 to 109 mg/kg-day, respectively. Exposure of the
parental generation (P) was initiated when the test animals were approximately 43 days of age and
continued through a 70-day pre-mating period and a cohabitation period. P generation males
were exposed for approximately 106 days prior to sacrifice. Exposure of P generation female rats
continued through gestation and lactation. F, generation rats were exposed to BDCM in utero
and by consumption of the dam's drinking water during the lactation period. At weaning, F, rats
(30/sex/concentration) were selected for a postweaning/premating exposure period, followed by a
cohabitation period, and the exposure continued through gestation and lactation. F, generation
females delivered litters and the F2 litters were sacrificed on lactation day 22.
Deaths at 150 and 450 ppm were associated with reduced water consumption, weight loss
and/or adverse clinical signs and may have been compound-related. Adverse clinical signs
occurred at 150 and 450 ppm were attributed to reduced water consumption. Body weight and
body weight gain were significantly reduced in the 450 ppm P generation males and females and
150 and 450 ppm F, generation males and females, and was associated with decreased food
consumption. Rats at 450 ppm also had decreased absolute and relative organ weights. Water
consumption was significantly reduced in P and Fx generation males and females at all
concentrations of BDCM, and was reduced by 10-20% at 150 and 450 ppm. There were no gross
pathological or histopathological indications of compound-related toxicity. Most indicators of
reproductive or developmental toxicity examined were not significantly affected by BDCM
treatment. However, decreased pup body weight was observed at weaning and/or during
lactation in the Fx and F2 generations in the 150 and 450 ppm groups. Small, but statistically
significant, delays in F, generation sexual maturation occurred at 150 (males) and 450 ppm
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(males and females) as determined by timing of vaginal patency or preputial separation, but no
statistically significant delays were seen when body weight at weaning was included as a covariate
in the analysis. Delayed estrus was observed in F, females in the 450 ppm exposure group. The
study authors also considered this effect to be a secondary response associated with reduced pup
weights. The results of this study appear to identify NOAEL and LOAEL values for reproductive
effects of 50 ppm (4.1 to 12.6 mg/kg-day) and 150 ppm (11.6 to 40.2 mg/kg-day), respectively,
based on delayed sexual maturation. However, the study authors have questioned whether
delayed sexual maturation in F, males associated with reduced body weight should be treated as
reproductive toxicity or general toxicity, since the root cause appears to be dehydration brought
about by taste aversion to the compound. The parental NOAEL and LOAEL are also 50 and 150
ppm, respectively, based on reduced body weight and body weight gain in F, males and females.
Klinefelter et al. (1995) evaluated the effects of BDCM exposure on male reproduction at
an interim sacrifice as part of a two-year bioassay, in which F344 rats were administered BDCM
in drinking water at concentrations of 0, 330 mg/L, or 620 mg/L. The authors estimated the
doses to be 0, 22, and 39 mg/kg/day. At 52 weeks, the authors conducted an interim sacrifice,
which included an evaluation of epididymal sperm motion parameters and histopathology of the
testes and epididymides. No histologic alterations were observed in any reproductive tissue.
Sperm velocities (mean straight-line, average path, and curvilinear), however, were significantly
decreased at 39 mg/kg/day. No effect on sperm motility was observed at 22 mg/kg/day. The
NOAEL and LOAEL for reproductive effects are thus 22 and 39 mg/kg/day, respectively.
Several studies on the developmental toxicity of BDCM gave negative results at doses up
to 116 mg/kg/day in Sprague-Dawley rats (NTP, 1998a) and 55.3 mg/kg/day in rabbits (CCC,
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2000c) when administered in drinking water. However, in other studies (CCC, 2000b; Ruddick et
al., 1983), slightly decreased numbers of ossification sites in the hindlimb and forelimb were
observed in fetuses of Sprague-Dawley rats administered 82 mg/kg/day in the drinking water on
gestation days 6 to 21 (CCC, 2000b) and sternebral aberrations were observed in the offspring of
Sprague-Dawley rats administered 200 mg/kg/day by gavage in corn oil on gestation days 6 to 15
(Ruddick et al., 1983). Reductions in mean pup weight gain and pup weight were observed when
parental Sprague-Dawley rats were administered BDCM in the drinking water at concentrations
of 150 ppm and above (a biologically single meaningful estimate of intake on a mg/kg/day basis
could not be calculated for this study) (CCC, 2000a). Full litter resorption has been noted in
F344 rats (Narotsky et al., 1997a; Bielmeier et al., 2001), but not Sprague-Dawley rats, treated
with BDCM at doses of 50 to 100 mg/kg/day during gestation days 6 to 10 or with a single dose
of 75 mg/kg/day on gestation day 9. Based on these observations, it appears that BDCM
administered in the drinking water can induced various developmental effects, and the types of
these effects are species- and strain-dependent.
Systemic Effects
BDCM causes decreased weight gain and various adverse effects in the nervous and
immune systems, thyroid, kidney, and liver, but the predominant systemic effects from acute and
chronic exposure to BDCM are on the liver and kidney. NTP (1987) administered BDCM to rats
by gavage in corn oil at doses of 0, 50 or 100 mg/kg/day for 102 weeks. Histologic alterations in
the liver and kidney were observed at 50 mg/kg/day and higher. In a similar study in mice (NTP,
1987), histologic alterations in the liver, kidney and thyroid of male mice were noted at 25
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mg/kg/day and higher doses. Aida et al. (1992) administered microencapsulated BDCM to rats in
the diet at dose levels ranging from 6 to 168 mg/kg/day for 24 months. At 6 mg/kg/day, liver
fatty degeneration and granuloma were observed.
Carcinogenicity
BDCM has been found to cause tumors in multiple target organs of multiple species.
Tumors were observed in the large intestine and kidneys of male and female rats, kidneys of male
mice and livers of female mice when these rodents were treated with BDCM in corn oil via
gavage in a 2-year bioassay (NTP, 1987). The NTP concluded that under the conditions of these
2-year gavage studies, clear evidence of carcinogenic activity existed for male and female rats and
mice. A study investigating the relationship between liver toxicity and tumorigenicity of BDCM
concluded that the shape of the dose-response curve was different for these two effects (Melnick
et al., 1998). The authors concluded that there does not appear to be a causal relationship
between liver toxicity and tumor development for BDCM.
In another rat study (George et al., 2002), BDCM induced heptocellular adenomas and
carcinomas in male rats exposed to the compound via drinking water for 104 weeks. However, a
biphasic dose-response was observed; the significant increase in hepatocellular neoplasmas only
occurred in the low dose group (8/45 vs. 2/45 in the control group) and not in the mid- and high-
dose groups (7/48 and 4/49, respectively). The underlying basis for the biphasic response is
unknown, but this pattern of response might be explained by inhibition of the hepatic metabolism
of BDCM by the compound itself. The same authors (George et al., 2002) also reported that
BDCM was not carcinogenic to male mice, but there is no evidence that an adequately high dose
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was tested in this study. DeAngelo (2002) reported mixed results on BDCM induction of
aberrant crypt foci (ACF), putative early preneoplastic lesions of the colon. The authors observed
increased ACF in a 13-week drinking water rat study, but not in 13- or 30-week mouse studies.
The study authors concluded that ACF induced by BDCM do not progress to neoplasia, as judged
by the absence of colon neoplasms in the two-year cancer study conducted by George et al.
(2002).
BDCM has shown mixed results in genotoxicity assays, with both positive (Fujie et al.,
1990; Simmon and Tardiff, 1978) and negative results (Hayashi et al., 1988; NTP, 1987; Varma
et al., 1988) from in vitro and in vivo studies. Synthesis of the overall weight of evidence from
these studies is complicated by the use of a variety of testing protocols, different strains of test
organisms, different activating systems, different dose levels, different exposure methods (gas
versus liquid) and, in some cases, problems due to evaporation of the test chemical. Although
data are mixed, the weight of evidence indicates that BDCM is genotoxic. Recent studies in
Salmonella strains engineered to contain rat GST-theta suggest that the mutagenicity of the
brominated THMs may be mediated by glutathione conjugation (DeMarini et al., 1997; Landi et
al., 1999). Furthermore, a recent study in female B6C3F1 mice (Melnick, et al., 1998) suggests
that increased incidence of hepatic tumors occurs at doses of BDCM that have no effect on
hepatocyte labeling index, indicating that regenerative hyperplasia is not required for tumor
induction, supporting the choice of linear low-dose extrapolation for quantification of cancer risk
associated with BDCM.
Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, BDCM is
classified as Group B2: Probable Human Carcinogen (EPA, 2002c). Under the EPA's 1999 Draft
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Guidelines for Carcinogen Risk Assessment, BDCM is likely to be carcinogenic to humans (EPA,
2002c). This descriptor is considered appropriate when there are no or inadequate data in
humans, but the combined experimental evidence demonstrates the production or anticipated
production of tumors by modes of action assumed to be relevant to humans. Based on the
evidence for BDCM genotoxicity, and the lack of data supporting a mode of action for which
nonlinear extrapolation to low doses are appropriate, a linear extrapolation to low doses is used
for quantification of the BDCM cancer risk.
Under the 1986 cancer guidelines, the Agency calculated a cancer oral slope factor, based
on renal tumors in treated male mice, of 6.2 x 10"2 per mg/kg/day and a unit risk of 1.8 x 10"6
(lig/L)"1 (EPA, 2002c). Based on the same endpoint, a cancer oral slope factor of 8.1 x 10"3 per
mg/kg/day has been calculated under the Draft Guidelines for Carcinogen Risk Assessment (EPA,
1999), corresponding to a unit risk of 2.3 x 10"7 (iig/L)"1 (EPA, 2002c). Based on this oral slope
factor, drinking water concentrations of 400 |ig/L, 40 |ig/L, and 4 |ig/L are estimated to be
associated with estimated lifetime risks of 10"4, 10"5, and 10"6.
Basis for RfD and MCLG
EPA selected the chronic study by Aida et al. (1992) as the most appropriate study for
derivation of the RfD. This study identified a LOAEL of 6 mg/kg/day based on liver fatty
degeneration and granuloma in male rats, as well as a BMD of 1.9 mg/kg/day and a corresponding
BMDL10 of 0.8 mg/kg/day, based on the same endpoint. The LOAEL of 6.1 mg/kg/day was used
as the basis for the RfD. An UF of 3000 was applied to the LOAEL: a 10-fold UF for
interspecies extrapolation; another 10-fold factor for protection of sensitive human
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subpopulations; a 10-fold factor for extrapolation from a LOAEL; and a 3-fold factor for database
deficiencies, including lack of a multigeneration reproductive toxicity study and uncertainty
related to possible reproductive or developmental effects suggested by epidemiological studies
(EPA, 2002c). These calculations result in an RfD of 0.002 mg/kg/day. The MCLG for BDCM is
zero, based on its probable human carcinogenicity, and linear low-dose extrapolation in light of
the evidence for a genotoxic mode of action, coupled with the absence of evidence for other
modes of action.
Children's Risk in Relation to the MCLG
The Agency believes that the MCLG of zero is protective for both the carcinogenic and
systemic effects of BDCM in children and adults. There is not sufficient evidence from studies on
the systemic effects of BDCM, or on the metabolism of the compound, to determine whether
children are more sensitive to the toxic effects of BDCM than are adults.
2.2.2.2. Dibromochloromethane
Developmental/Reproductive Effects
In a prospective study, Windham et al. (2003) demonstrated that increasing levels of
individual brominated THMs or total brominated THMs in the drinking water were associated
with significantly shorter cycles when examined by quartile. Similar decrements were observed in
follicular, but not luteal, phase length. The strongest association for an individual compound was
observed for DBCM with adjusted decrements of 1.2 days (95% C.I. -2.0, -0.38) for mean cycle
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length and 1.1 days (95% C.I. -1.9, -0.25) for mean follicular phase length at the highest quartile
(>20 ng/L).
Ruddick et al. (1983) investigated developmental toxicity in pregnant Sprague-Dawley
rats (9-15/group) administered DBCM by gavage in corn oil at dose levels of 0, 50, 100, or 200
mg/kg/day from gestation days 6 to 15. Although maternal toxicity was indicated by a significant
decrease in body weight gain at the highest dose, no treatment-related developmental toxicity was
observed for DBCM. No statistical analysis was performed in this study and inspection of the
data revealed no dose-related effects. The power of the experiment was limited by the small
number of litters per dose group. Thus, the developmental NOAEL was 200 mg/kg/day and a
LOAEL could not be identified. Borzelleca and Carchman (1982) conducted a two-generation
reproductive study in ICR Swiss mice. Nine-week-old mice (10 males and 30 females per dose
group) were continuously maintained on drinking water containing 0, 100, 1000, or 4000 mg/L
DBCM (0, 17, 171, or 685 mg/kg/day). Based on postnatal body weight in the F2b pups, this
study identified a marginal LOAEL of 17 mg/kg/day for DBCM, and a NOAEL could not be
determined. This LOAEL was considered to be minimal because the body weight decrease was
only observed at one time point, the effect was only noted in one of the two litters in the F2
generation, no other adverse effects were noted at the dose level, and it was unclear from the
report how many litters and pups per litter were examined for postnatal body weight.
NTP (1996) conducted a short-term reproductive toxicity screening study on Sprague-
Dawley male and female rats. A group of male rats and two groups of female rats were treated
with DBCM in drinking water at concentrations of 0, 50, 150, or 450 mg/L (10 rats/dose/group)
during a study period of 35 days (from gestation day 6 through parturition). Males were treated
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on study days 6-34. Group A females were treated on study days 1-34, and were mated to
treated males on study days 13-18. Group B females were mated on study day 1 to untreated
males and treated from gestation day 6 through parturition. Based on measured water
consumption, the authors estimated dose levels for the treated males to be 0, 4.2, 12.4, or 28.2
mg/kg/day, and for treated females to be 6.3, 17.4, or 46.0 mg/kg/day (Group A) and 7.1, 20.0,
or 47.8 mg/kg/day (Group B). The developmental toxicity of the offspring from these treated rats
was compared with those from the control group. No significant reproductive/ developmental
toxicity was observed at any dose level. The NOAEL for reproductive/ developmental effects
identified in this study was 28.2 mg/kg/day for males and 47.8 mg/kg/day for females; no LOAEL
was identified. The study sensitivity was decreased by the use of a relatively small number of
animals per group and the lack of microscopic examination of the pups.
Systemic Effects
DBCM causes decreased weight gain and various adverse effects in the nervous and
immune systems, kidneys, and liver. The predominant effects from acute and chronic exposure to
DBCM are on the liver and kidney. Tobe et al. (1982) administered microencapsulated DBCM in
the diet to rats for 24 months and reported decreased body weight and changes in clinical
chemistry parameters and gross liver appearance in males at 49 mg/kg/day; similar effects were
seen in females at slightly higher administered doses. Tobe et al. (1982) identified a NOAEL of
12 mg/kg/day, but this NOAEL is uncertain in light of the lack of histopathological examination
and the results of Aida et al. (1992) with BDCM, finding that adverse liver histopathology occurs
at doses lower than those observed by Tobe et al. (1982) based on clinical chemistry and gross
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tissue examination. NTP (1985) reported histologic lesions in the liver in male and female rats at
40 mg/kg/day and lesions in the liver and thyroid in female mice at 50 mg/kg/day after exposure
to DBCM via gavage in oil for 104 weeks; no NOAELs were identified. NTP (1985) evaluated
the toxicologic effects of DBCM after a subchronic (13 week) exposure. Hepatic lesions were
identified at 60 mg/kg/day, with a subchronic NOAEL 30. Adjusting for exposure 5 days/week,
the subchronic NOAEL was 21.4 mg/kg/day, and the LOAEL was 43 mg/kg/day.
Carcinogenicity
Evidence is limited for DBCM carcinogenicity via ingestion. An increase in tumors
occurred in the livers of male and female mice, while no increase in tumors occurred in male and
female rats, when these rodents were treated with DBCM via gavage in corn oil for 2 years (NTP,
1985). In the mouse study (NTP, 1985), a significant increase in hepatocellular carcinomas and
adenomas occurred in female high-dose group (19/50 vs. 6/50). The combined incidence of
hepatocellular adenomas or carcinomas in high dose male mice (27/50 vs. 23/50) was also
significant, but this effect is considered marginal overall because it was only significant in the life
table test but was not significant (p=0.065) by the incidental tumor test. The low-dose male
mouse group in this study was considered to be unsuitable for data analysis because an overdose
killed most of the male mice. No significant increase in carcinomas and adenomas was observed
in the female low-dose group (10/50 vs. 6/50). Based on these data, the NTP concluded that
there was equivocal evidence of DBCM carcinogenicity in male B6C3F1 mice, some evidence of
carcinogenicity in female B6C3F1 mice, and no evidence of carcinogenicity in male or female rats.
Based on a comparison of the shape of the dose-response curve for liver toxicity and
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tumorigenicity of DBCM, Melnick et al. (1998) concluded that there does not appear to be a
causal relationship between liver toxicity and tumor development for DBCM. De Angelo (2002)
reported that in a 13-week drinking water study, DBCM can induce aberrant crypt foci (ACF),
putative early preneoplastic lesions, in the colons of male rats. However, it is not clear whether
this change would result in colon cancer in exposed rats.
DBCM has shown mixed results in genotoxicity assays, with both positive (e.g., Sekihashi
et al., 2002; Simmon and Tardiff, 1978; Sofuni et al., 1996) and negative results (e.g., Hayashi et
al., 1988; NTP, 1985; Potter et al., 1996) from in vitro and in vivo studies. Synthesis of the
overall weight of evidence from these studies is complicated by the use of a variety of testing
protocols, different strains of test organisms, different activating systems, different dose levels,
different exposure methods (gas versus liquid) and, in some cases, problems due to evaporation of
the test chemical. EPA (1994b) has previously determined that the weight of evidence for DBCM
mutagenicity and genotoxicity is inconclusive. Recent studies in Salmonella strains engineered to
contain rat GST-theta suggest that the mutagenicity of the brominated THMs may be mediated by
glutathione conjugation (DeMarini et al., 1997).
Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, DBCM is
classified as Group C: Possible Human Carcinogen, based on inadequate human data and limited
evidence of carcinogenicity in animals (EPA, 2002c). Based on the 1999 Draft Guidelines for
Cancer Risk Assessment, there is suggestive evidence of human carcinogenicity of DBCM, but
the data are not sufficient to assess human carcinogenic potential. A compound is described as
having suggestive evidence of carcinogenicity when the evidence from human or animal data is
suggestive of carcinogenicity, but is judged not sufficient for a conclusion as to human
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carcinogenic potential (e.g., a marginal increase in tumors that may be exposure-related, or
evidence is observed only in a single study). The evidence for DBCM carcinogenicity is weaker
than the evidence for the other brominated THMs. However, the evidence for DBCM
carcinogenicity is strengthened by the evidence for carcinogenicity of BDCM and bromoform,
both of which are considered likely to be carcinogenic to humans. BDCM, DBCM, and
bromoform are all brominated THMs with increasing numbers of bromine atoms, suggesting that
DBCM is likely to share toxic effects common to both BDCM and bromoform.
Under the 1986 Guidelines for Carcinogen Risk Assessment, the Agency calculated a
cancer oral slope factor of 8.4 x 10"2 per mg/kg/day and a unit risk of 2.4 x 10"6 (|ig/L)"' based on
liver tumors in female mice dosed via oral gavage in corn oil (EPA, 2002c). Dose-response
assessment is not recommended under the 1999 guidelines for chemicals for which the weight of
evidence is described as "suggestive evidence of human carcinogenicity, but not sufficient to
assess human carcinogenic potential
Basis for RfD and MCLG
EPA selected the NTP subchronic study in rats (NTP, 1985) as the most appropriate basis
for derivation of the RfD and DWEL. This study identified a NOAEL of 30 mg/kg/day (duration
adjusted to 21.4 mg/kg/day). An UF of 1000 was applied to the subchronic NOAEL: a 10-fold
factor for interspecies extrapolation; a 10-fold factor for protection of sensitive human
subpopulations; and a 10-fold factor for subchronic to chronic extrapolation (EPA, 2002c). The
resulting RfD was 0.02 mg/kg/day, corresponding to a DWEL of 0.7 mg/L for a 70 kg adult
drinking 2 L of water/day.
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A relative source contribution (RSC) of 80% is used for DBCM, based on an analysis of
the anticipated potential for exposure to DBCM in disinfected tap water via ingestion, inhalation,
and dermal contact, and because exposure via other media (outdoor air, food, and soil) are
anticipated to be low (EPA, 2002c). There are some uncertainties in the 80% RSC that are
related to the availability of adequate concentration data for DBCM in media other than water.
The MCLG of 0.06 mg/L for DBCM is based on a noncarcinogenic endpoint (the RfD)
with an additional safety factor of 10 to account for possible carcinogenicity. Thus, the MCLG is
calculated as: MCLG = (0.02 mg/kg/day x 70 kg x 0.8)/(2 L/day x 10) = 0.056 mg/L, rounded
to 0.06 mg/L.
Children's Risk in Relation to the MCLG
Developmental and reproductive toxicity data are available for DBCM and were
considered in the derivation of the RfD for DBCM. The NOAEL in the two-generation
reproductive study of Borzelleca and Carchman (1982) (17 mg/kg/day) is comparable to the
duration-adjusted NOAEL (21.4 mg/kg/day) that was the basis for the RfD. There is not
sufficient evidence from studies on the systemic effects of DBCM, or on the metabolism of the
compound, to determine whether children are more sensitive to the toxic effects of DBCM than
are adults. Nevertheless, the Agency believes that the MCLG of 0.06 mg/L is protective of
children's health, because developmental or reproductive effects have not been found to occur
below the level of the critical effect (liver toxicity) used to derive the current RfD, and because
development of the MCLG includes the standard UF of 10 for protection of sensitive populations.
There is, however, some uncertainty in this conclusion.
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2.2.2.3. Bromoform
Developmental/Reproductive Effects
In a prospective study, Windham et al. (2003) demonstrated that increasing levels of
individual brominated THMs or total brominated THMs in the drinking water were associated
with significantly shorter cycles when examined by quartile. Similar decrements were observed in
follicular, but not luteal, phase length. For bromoform, the adjusted decrements were 0.79 days
(95% C.I. -1.4, -0.14) for mean cycle length and 0.78 days (95% C.I. -1.4, -0.14) for mean
follicular phase length at the highest quartile (>12 //g/L). The strongest association for an
individual compound was observed for DBCM.
Ruddick et al. (1983) investigated developmental toxicity in pregnant Sprague-Dawley
rats (9 -15/group) administered bromoform by gavage in corn oil at dose levels of 0, 50, 100, or
200 mg/kg/day from gestation days 6-15. No maternal toxicity was observed at any dose level;
some fetal skeletal anomalies were observed. Incidences of both fetuses and litters with
interparietal deviations were increased at the mid- and high-dose groups compared with the
controls. Furthermore, incidences of both fetuses and litters with sternebra aberrations increased
in a dose-related fashion. No statistical analysis was performed by the authors, but an
independent statistical analysis (EPA, 2002c, using the Fisher Exact test) demonstrated that the
increase in sternebral anomalies was significantly different from controls at 200 mg/kg/day. A
trend test showed a statistically significant dose-related trend (p<0.002) for this endpoint;
stepwise analysis indicated that this trend was no longer significant when the two highest doses
(100 mg/kg/day and 200 mg/kg/day) were removed from the analysis. These findings suggest that
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the NOAEL and LOAEL for developmental toxicity from this study were 50 mg/kg/day and 100
mg/kg/day, respectively. The NOAEL for maternal toxicity was 200 mg/kg/day, and a maternal
LOAEL could not be determined.
The prenatal and postnatal effects of bromoform on fertility and reproduction were
investigated by NTP (1989a) in Swiss CD-I mice using a continuous reproductive breeding
protocol. Mice were administered bromoform by gavage in corn oil at dose levels of 0, 50, 100,
or 200 mg/kg/day for 105 days, including a seven-day pre-cohabitation phase and a 98-day
cohabitation phase. Fx litters in the control and 200 mg/kg/day groups were raised to sexual
maturity (approximately 74 days) while receiving the same treatment as their parents. At sexual
maturity, males and females from different litters within the same treatment group were cohabited
for seven days and then housed individually until delivery. In the 200 mg/kg/day group, the body
weights of the dams at delivery were consistently less than the controls. No effect on any fertility
or reproductive parameter (numbers of litters per pair, litter size, proportion of live pups, sex ratio
of live pups, and pup body weight) was observed in the P generation. However, postnatal
survival of the Fx pups in the 200 mg/kg/day group was significantly less than in the control
group. As for the P generation, there was no effect on any mating, fertility or reproductive
parameter in the F, generation. At sacrifice, male and female F, mice administered 200
mg/kg/day exhibited increased relative liver weights and decreased relative kidney weights
compared to controls. Histopathological examination revealed minimal to moderate
hepatocellular degeneration in the livers of the 200 mg/kg/day male and female mice. Therefore,
based on liver histopathology, decreased postnatal survival, and other signs of toxicity (e.g.,
increased relative liver and decreased relative kidney weights) the developmental NOAEL and
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LOAEL from this study were 100 mg/kg/day and 200 mg/kg/day, respectively. The NOAEL and
LOAEL for maternal toxicity were also 100 mg/kg/day and 200 mg/kg/day, respectively, based on
consistently decreased body weights of pregnant dams at delivery.
Systemic Effects
Bromoform causes decreased weight gain and various adverse effects in the nervous
system, kidney, liver, and thyroid, but the predominant effects from acute and chronic exposure to
bromoform are on the liver. Tobe et al. (1982) administered bromoform microencapsulated in the
diet of rats for 24 months and reported gross liver lesions and changes in clinical chemistry
parameters in male rats at 90 mg/kg/day, but this study was not appropriate as the basis for the
RfD, because there was no histopathological analysis. NTP (1989b) conducted a chronic oral
study in rats and mice (50/sex/group) administered bromoform at 0, 100, or 200 mg/kg/day by
gavage in corn oil. This study observed histologic lesions in the liver (fatty changes and chronic
inflammation) at 100 mg/kg/day in male and female rats, and fatty changes in the liver of female
mice at the same dose. Female mice exposed to 200 mg/kg/day also exhibited follicular cell
hyperplasia of the thyroid gland. This study identified a LOAEL of 100 mg/kg/day based on liver
effects in both male and female rats and in female mice; no NOAEL was identified. For male mice,
a NOAEL of 100 mg/kg/day was identified.
NTP (1989b) also conducted a subchronic oral study in rats and mice. Male and female
F344/N rats (10/sex/dose) received 0, 12, 25, 50, 100, or 200 mg/kg/day bromoform by gavage in
oil, 5 days/week for 13 weeks. Male and female B6C3F1 mice received 0, 25, 50, 100, 200, or
400 mg/kg/day bromoform, for the same length of time as the rats. In male rats, a dose-
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dependent increase in the frequency of hepatocellular vacuolation was observed, which reached
statistical significance at 50 mg/kg/day. This effect was not observed in female rats. In male
mice, a dose-dependent increase in the number of hepatocellular vacuoles was seen at 200
mg/kg/day. Thus, this study identified a NOAEL and LOAEL of 25 mg/kg/day and 50
mg/kg/day, respectively, based on hepatocellular vacuolation in male rats, and a NOAEL and
LOAEL of 100 mg/kg/day and 200 mg/kg/day, respectively, based on hepatocellular vacuoles in
male mice.
Carcinogenicity
An increase in tumors occurred in the large intestine in male and female rats, while no
statistically significant increase in the incidence of tumors was seen in male or female mice, when
these rodents were treated with bromoform via gavage in corn oil (NTP, 1989b). The NTP
concluded that there was some evidence for carcinogenic activity in male rats, clear evidence in
female rats, and no evidence in male or female mice. De Angelo (2002) reported that in a 13-
week drinking water study, bromoform can induce ACF, putative early preneoplastic lesions, in
the colons of male rats. However, it is not clear whether this change would result in colon cancer
in exposed rats. In a study in which bromoform was administered by intraperitoneal injection in
mice, the number of lung tumors per mouse in the mid-dose group was significantly elevated over
controls (Theiss et al., 1977).
Bromoform has shown mixed results in mutagenicity assays, with both positive (Simmon
and Tardiff, 1978; Zeiger, 1990) and negative (Hayashi et al., 1988; Ishidate et al., 1982; NTP,
1989a) results from in vitro and in vivo studies. Synthesis of the overall weight of evidence from
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these studies is complicated by the use of a variety of testing protocols, different strains of test
organisms, different activating systems, different dose levels, different exposure methods (gas
versus liquid) and, in some cases, problems due to evaporation of the test chemical. Although
data are mixed, the weight of evidence indicates that bromoform is genotoxic. In addition, recent
studies in Salmonella strains engineered to contain rat GST-theta suggests that the mutagenicity
of the brominated THMs may be mediated by glutathione conjugation (De Marini et al., 1997;
Landi et al., 1999).
The evidence is sufficient to consider bromoform a Group B2: Probable Human
Carcinogen via ingestion under the EPA's 1986 Guidelines for Cancer Risk Assessment (EPA,
2002c). Based on the observed rare large intestinal tumors observed in both sexes of rats tested
and positive genotoxicity results, bromoform can be described under the 1999 Draft Guidelines
for Carcinogen Risk Assessment as likely to be carcinogenic to humans (EPA, 2002c). Based on
the evidence for bromoform genotoxicity, and the lack of data supporting a mode of action for
which nonlinear extrapolation to low doses is appropriate, a linear extrapolation to low doses is
used to quantify the bromoform cancer risk.
The Agency calculated a cancer oral slope factor of 7.9 x 10"3 per mg/kg/day and a unit
risk of 2.3 x 10"7 per (|ig/L)(EPA, 2002c), based on the incidence of intestinal tumors in rats, as
specified in the 1986 Guidelines for Carcinogen Risk Assessment (EPA, 1986). Calculating a
cancer slope factor based on the Draft Guidelines for Carcinogen Risk Assessment (EPA, 1999),
using the same endpoint results in a value of 4.5 xlO"3 (mg/kg/day)"1, corresponding to a unit risk
of 1.3 x 10"7 (|ig/L)"' (EPA, 2002c). Based on this calculation, drinking water concentrations of
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approximately 800 |ig/L, 80 |ig/L and 8 |ig/L are estimated to be associated with lifetime cancer
risks of 10"4, 10"5 and 10"6, respectively.
Basis for RfD and MCLG
EPA derived the RfD for bromoform based on hepatocellular vacuolization in the liver of
male rats in the NTP (1989b) subchronic study. This study identified a NOAEL of 25 mg/kg/day
and a LOAEL of 50 mg/kg/day. After adjustment for dosing 5 days/week, the duration-adjusted
NOAEL and LOAEL were 17.9 and 35.7 mg/kg/day, respectively. The duration-adjusted
BMDL10 was 2.6 mg/kg/day. The subchronic study was preferred to the chronic study as the
basis for the RfD. This is because there was much less uncertainty in the estimate of the BMDL10
for the subchronic study, due to the doses in the subchronic study being closer together, and in
light of the high response at the lowest dose of the chronic study. Furthermore, a NOAEL was
not identified in the chronic study, and the critical effect was consistent across both the subchronic
and chronic studies. Thus, the RfD was based on a NOAEL of 17.9 mg/kg/day and a composite
UF of 1000: 10-fold factors each for interspecies variation and for protection of sensitive
subpopulations; and a 10-fold factor for subchronic to chronic extrapolation. The database for
bromoform includes systemic toxicity studies in two species, a two-generation study in mice, and
a developmental toxicity study in rats. In light of the other data on the THMs, for which the
critical effect is systemic toxicity, the single data gap of a developmental toxicity study in a second
species is insufficient to require a database UF. The resulting RfD is 0.02 mg/kg/day.
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The MCLG for bromoform is zero, based on its probable human carcinogenicity, and
linear low-dose extrapolation in light of the evidence for a genotoxic mode of action, coupled
with the absence of evidence for other modes of action.
Children's Risk in Relation to the MCLG
The Agency believes that the MCLG of zero is protective for both the carcinogenic and
systemic effects of bromoform in children and adults. There is no evidence from studies on the
systemic effects of bromoform, or on the metabolism of the compound, that children are more
sensitive to the toxic effects of bromoform than are adults.
2.3. HALOACETIC ACIDS
There are only limited data available for determining children's risk resulting from
exposure to haloacetic acids (HAAs). The haloacetic acids have been tested to varying degrees in
developmental and reproductive toxicity studies in animals, as described below for each HAA
covered in this document. Human epidemiology data on the developmental and reproductive
toxicity of the haloacetic acids are lacking. Most of the human health data for halogenated acetic
acids are as components of complex mixtures of water disinfection byproducts. Although most
studies of human health effects following exposure to water disinfection byproducts have used
total trihalomethanes as the exposure metric, Klotz and Pyrch (1999), conducted a case-control
study on the relationship between neural tube defects and drinking water exposure to
trihalomethanes, haloacetonitriles or haloacetic acids. (See Section 2.1 for more study details.) A
statistically significant prevalence odds ratio (POR) was reported for the highest tertile (third) of
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trihalomethane exposure; however, only a slight non-statistically significant excess risk (POR 1.2:
95% confidence interval 0.5-2.6) was found for cases when analyzed based on total haloacetic
acids tertiles. The specific haloacetic acids that were measured as part of the total haloacetic acid
exposure estimate were not identified. Based on the results of the study, the authors concluded
that there was no clear association of HAAs with neural tube defects.
Only a subset of haloacetic acids has been tested in standard cancer bioassays. The
chlorinated acetic acids have been tested, but published reports of full cancer bioassays are not yet
available for their brominated counterparts. Cancer studies in animals are described separately for
each compound below.
The data are insufficient to determine whether there are age-dependent differences in the
metabolism of the haloacetic acids that might lead to differences in health risk. The enzymes
responsible for the metabolism of monochloroacetic acid (MCA) and monobromoacetic acid
(MBA) have not been identified. The metabolism of the dihaloacetic acids has been more
extensively studied. Dichloroacetic acid (DCA), bromochloroacetic acid (BCA) and
dibromoacetic acid (DBA) all are metabolized to glyoxylic acid through a reaction catalyzed by
glutathione-S-transferase-Zeta (GST-Zeta) (Tong et al., 1998a; Tong et al., 1998b). Glyoxylic
acid is in turn metabolized through a variety of competing pathways to form glycine, glycolate,
oxalate, or C02 (Stacpoole et al., 1998). Since DCA is a potential metabolite of trichloroacetic
acid (TCA) (Bull 2000; Lash et al., 2000), this enzyme pathway may also be relevant for
evaluating the susceptibility of children to TCA. Even in the cases where relevant metabolizing
enzymes have been identified, there is no information on age-dependent changes in the expression
or activity of these enzymes. It is noteworthy, however, that Stacpoole et al. (1998) reported that
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the toxicokinetics of DC A is similar in children and adults given pharmacological doses of this
compound, suggesting that there are no significant age-dependent differences in DCA metabolism.
The health implications of any differences between children and adults in metabolic capacity are
also difficult to determine for the haloacetic acids; neither the mode of action nor toxic forms are
known.
2.3.1. Monochloroacetic Acid
Developmental/Reproductive Effects
Johnson et al. (1998) studied the developmental toxicity of MCA by exposing pregnant
Sprague-Dawley rats on gestation days 1-22 to 0 or 1570 mg/L MCA in drinking water (reported
by the authors to be equivalent to doses of 0 or 193 mg/kg/day). Dams were sacrificed on
gestation day 22, and implantation sites, resorption sites, fetal placements, fetal weights, placental
weights, crown-rump lengths, gross fetal abnormalities, and abnormal abdominal organs were
recorded. In addition, the hearts were evaluated microscopically for abnormalities. There was no
microscopic evaluation of other internal anomalies, and no evaluation of skeletal anomalies. All
dams survived without evidence of toxicity. Although the authors reported that "weight gain
during pregnancy was not significantly different for any group of dams," the average maternal
weight gain in the exposed rats was markedly reduced (18 g/dam, compared to 122.1 g/dam in the
controls). The magnitude of the decreased weight gain was confirmed by the study authors
(Johnson, personal communication), but it remains unclear why this difference was not considered
statistically significant. No adverse effects on reproductive or developmental endpoints were
reported. There was no effect on any other evaluated endpoint of maternal or fetal toxicity.
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Based on the marked decrease in weight gain, the single dose tested of 193 mg/kg/day is a
maternal LOAEL and a developmental NOAEL. There was no evidence to suggest any
developmental toxicity resulting from exposure to MCA. This lack of effect occurred despite the
apparently large differences in maternal weight gain during exposure. Complete fetal
examinations were not conducted, and the study is limited by the small size of the exposed group.
Smith et al. (1990 in a published abstract) reported on the results of a developmental
toxicity study in which pregnant Long-Evans rats were treated with MCA at doses of 0, 17, 35,
70, or 140 mg/kg/day by gavage on gestation days 6-15. MCA treatment significantly reduced
maternal weight gain at 140 mg/kg/day but did not affect maternal deaths or organ weight at any
doses. No treatment-related effects on resorptions/litter or pup weight were observed. The mean
frequency per litter of soft tissue malformations ranged from 1.2% in the control group to 6.4% in
the high-dose group (140 mg/kg/day); the change was not considered to be dose-related. The
highest dose of MCA caused a significantly increased incidence of malformations of the
cardiovascular system, mainly levocardia (primarily a defect between the ascending aorta and right
ventricle). There were no skeletal malformations observed in this study. Based on the
malformations of the cardiovascular system, the LOAEL for developmental toxicity was 140
mg/kg/day, and the next lower dose of 70 mg/kg/day can be considered a NOAEL. Maternal
toxicity was observed at the LOAEL for developmental toxicity, as evidenced by a significant
reduction in maternal weight gain. Complete details of the study methods and results were not
published.
MCA has been evaluated in in vitro systems for developmental toxicity. Hunter et al.
(1996) treated early somite-staged conceptuses (3-6 somites) from CD-I mice to MCA
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concentrations ranging from 0 to 500 |iM. Statistically significant, concentration-dependent
increases in malformations were seen at sub-lethal doses. MCA has also been tested in
developmental screening assays in non-mammalian systems. The Hydra system is an assay that
determines the degree to which a test chemical can perturb embryonic development at maternally
subtoxic doses. It has been purposely designed to overestimate developmental hazard potential
and its primary utility is as a screening tool for identifying compounds for in vivo developmental
toxicity testing (Fu et al., 1990). This assay system compares the toxicity to adult Hydra and to
artificial "embryos" made from disassociated and randomly reaggregated terminally differentiated
and pluripotent stem cells. Using this system, Fu et al. (1990) studied MCA's developmental
toxicity. Based on the ratio of minimal effective toxic concentrations for adults and artificial
embryos, they reported that MCA was more than 8 times more toxic to in vitro development than
to the adult component of the assay. Ji et al. (1998) also studied the teratogenic potential of
MCA using Hydra. The teratogenic potential was estimated by the ratio of toxicity to
regenerative inhibition (T50/I50), which yielded a ratio of 6.16. The study authors concluded that
MCA had a high teratogenic potential. The in vitro studies using mammalian whole-embryo
culture and Hydra provide support for the developmental toxicity of MCA observed in vivo
(Smith et al., 1990).
No studies were located on the reproductive toxicity of MCA.
Systemic Effects
Effects in several short-term MCA toxicity studies included neurotoxicity at high doses
near the LD50. In lower-dose studies, increased nasal discharge and lacrimation were observed in
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rats beginning at 7.5 mg/kg/day, but no adverse effects were observed in mice treated at similar
doses (NTP, 1992b). Subchronic and chronic studies suggest that the primary targets for MCA-
induced toxicity include the heart, spleen and nasal epithelium. In a 13-week oral gavage study in
rats, decreased heart weight was observed at 30 mg/kg/day and cardiac lesions progressed in
severity with increasing dose. Liver and kidney toxicity were only observed at higher doses
(NTP, 1992b). In 2-year oral gavage studies, decreased survival was observed at 15 mg/kg/day in
rats and decreased survival and nasal and forestomach hyperplasia were observed in mice at 100
mg/kg/day (NTP, 1992b). A more recent study appears to confirm these findings (DeAngelo et
al., 1997); this study identified a critical effect of increased spleen weight.
Carcinogenicity
MCA did not induce a carcinogenic response in two chronic rodent bioassays (NTP,
1992b; DeAngelo et al., 1997). There was no evidence for carcinogenicity in the NTP (1992b)
study; however, the route of compound administration was via gavage, only two doses were
tested, and significant mortality was observed in high-dose male rats, high-dose male mice, and
low- and high-dose female rats. The high mortality rates may have compromised the power and
sensitivity of the study to detect MCA-associated tumor effects.
In the DeAngelo et al. (1997) drinking water study, only male F344 rats were tested.
Female F344 rats may be more sensitive to MCA-associated effects, as evidenced by some of the
toxicity findings in the subchronic oral gavage study (NTP, 1992b; Bryant et al., 1992). MCA has
not been tested for carcinogenicity in a drinking water assay in a second species.
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MCA has yielded mixed results in genotoxicity assays. MCA was negative in the S.
typhimurium reverse mutation assay (Mortelmans et al., 1986) but positive in the mouse
lymphoma gene mutation assay in the absence of metabolic activation (McGregor et al., 1987). It
was positive for the induction of sister chromatid exchanges in Chinese hamster ovary cells in the
absence of S9 (liver enzyme preparation containing cytochrome P450) activation (but not in the
presence of S9) (Galloway et al., 1987), but was negative for sister chromatid exchanges in a
Chinese hamster lung fibroblast system (Sawada et al., 1987). MCA did not induce strand breaks
in rats or mice (Chang et al., 1991). In a more recent study including gene mutation assays in
Escherichia coli and Salmonella TA100 and the newt (Pleurodeles waltl larvae) micronucleus
test, MCA gave uniformly negative results (Giller et. al, 1997).
Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment (EPA, 1986),
MCA is best classified as Group D: Not Classifiable as to Human Carcinogenicity. According to
the 1999 Draft Guidelines for Carcinogen Risk Assessment (EPA, 1999), the data on MCA are
inadequate for an assessment of human carcinogenic potential.
Basis for RfD and MCLG
The Agency derived the MCA RfD from the LOAEL of 3.5 mg/kg/day for increased
spleen weight in the chronic rat drinking-water study (DeAngelo et al., 1997). A composite UF
of 1000 was used: a 10-fold factor for interspecies extrapolation, a 10-fold factor for human
variability, a 3-fold factor for extrapolation from a minimal LOAEL, and a 3-fold factor for
database deficiencies, including lack of adequate developmental toxicity studies in two species and
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the lack of a multi-generation reproductive study. The resulting RfD is 0.004 mg/kg/day,
corresponding to a DWEL of 0.14 mg/L for a 70 kg adult drinking 2 L of water/day.
EPA calculated the MCLG by applying a relative source contribution (RSC) of 20% to the
DWEL: MCLG = 0.14 mg/L x 0.2 = 0.028 mg/L (rounded to 0.03 mg/L). The resulting
proposed MCLG for MCA is 0.03 mg/L. The default RSC of 20% was chosen in accordance
with the exposure decision tree approach in EPA's Human Health Methodology (EPA, 2000f),
taking into account the likelihood of exposure to MCA from sources other than tap water, such as
ambient air and food. The available data are sufficient to demonstrate that food and air are
relevant exposure sources in addition to drinking water, but the data are inadequate to calculate
an RSC.
Children's Risk in Relation to the MCLG
There is no evidence from studies on the systemic effects of MCA that children are more
sensitive to the toxic effects of MCA than are adults. However, these data are limited. No data
on potential metabolic differences between children and adults for MCA were located.
The only developmental effects that have been reported for MCA are malformations of the
cardiovascular system reported in a published abstract (Smith et al., 1990). The NOAEL was
70 mg/kg/day, and the LOAEL for developmental toxicity was 140 mg/kg/day, where
maternal toxicity also occurred (Smith et al., 1990). In contrast, Johnson et al. (1998) reported
that the single dose tested of 193 mg/kg/day induced maternal, but not developmental effects.
Taken together, these data suggest that the developing fetus is no more sensitive to MCA than are
adults. The LOAEL for systemic toxicity of 3.5 mg/kg/day (DeAngelo et al., 1997) is 20x lower
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than the NOAEL for developmental effects of 70 mg/kg/day. The MCLG of 0.03 mg/L for
systemic effects is adequately protective of fetuses and children.
2.3.2. Dichloroacetic Acid
Developmental/Reproductive Effects
In two studies reported by Smith et al. (1992), pregnant Long-Evans rats (approximately
20/dose) received DCA by gavage on gestation days 6-15 at doses of 0, 900, 1400, 1900, or 2400
mg/kg/day (first study) or 0, 14, 140, or 400 mg/kg/day (second study). Dams were sacrificed on
gestation day 20, and both maternal toxicity and fetal toxicity were assessed. Dose-related
increases in mortality occurred in dams dosed at 1400 mg/kg/day and above, body weight gain
was significantly reduced at 140 mg/kg/day and above, and increased liver weight was observed in
all the treated groups. Significant implantation loss occurred at 900 mg/kg/day and above, and
the number of live fetuses per litter was reduced at 900 mg/kg/day and above. Fetal weight and
crown-rump length were significantly lower at levels of 400 mg/kg/day and above. Dose-related
increases were also reported for external, soft tissue, cardiovascular, urogenital, and orbital
malformations in the developing fetuses at doses of 140 and above. The authors identified a
developmental NOAEL of 14 mg/kg/day and a LOAEL of 140 mg/kg/day. Based on increased
liver weight, the maternal NOAEL and LOAEL were also 14 and 140 mg/kg/day, respectively.
Epstein et al. (1992) reported findings from a series of experiments in pregnant Long-
Evans rats exposed to DCA by gavage. There were three separate, sequential phases of this
study; in each phase, the dams were exposed for a specific 1- to 3-day period during gestation and
were sacrificed on gestation day 20. Both maternal and fetal toxicity were assessed, including
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histological examination of the fetuses. In all three phases of the study, no treatment-related
maternal toxicity was observed (based on body weight and organ weight data). In the first phase
of the study, dams were exposed to 1900 mg/kg/day during gestation days 6-8, 9-11, or 12-15, in
order to observe the effects of DCA during specific periods of organogenesis. A decrease in
average fetal weight was reported in the dose group exposed on days 6 to 8, but no
malformations were reported. In the groups dosed on days 9-11 and 12-15, the mean percentage
of cardiac malformations per litter was significantly (p<0.001) increased. In the second phase of
the study, pregnant dams were administered a single dose of 2400 mg/kg on gestation days 10,
11, 12, or 13. Fetal weights in each exposure group were similar to control values. Significant
(p<0.05) increases in cardiac malformations were reported in groups exposed on day 10 or day
12. In the third phase of the study, a single dose of 3500 mg/kg was administered to dams on
gestation days 9, 10, 11, 12, or 13. This higher dose level, 3500 mg/kg, resulted in a slightly
higher incidence of cardiac defects (3.6% vs. 2.9% in controls), and the increase was significant
(p<0.05) on day 12. The results from these studies suggest that acute high-dose treatments of
DCA at specific developmental stages can induce developmental toxicity in the absence of
maternal toxicity.
The developmental toxicity of DCA has also been evaluated in in vitro systems. Saillenfait
et al. (1995) exposed explanted embryos from Sprague-Dawley rats for 46 hours to 0, 1.0, 2.5,
3.5, 5.0, 7.5, or 10 mM DCA. A significant, dose-dependent decrease in crown-rump length was
seen at 3.5 mM and above. In addition, several effects (brain and eye defects, reduced embryonic
axis) were seen at 2.5 mM. A similar study with CD-I mouse whole-embryo culture exposed to
DCA for 24 hours found significant increases in neural-tube defects at treatment concentrations of
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5.9 mM and above (Hunter et al., 1996). Another whole- embryo culture study was designed to
test mechanisms of haloacetic acid-induced dysmorphogenesis (Ward et al., 2000). The authors
reported that DCA treatment increased the accumulation of sub-Gl events (a measure of cells
with less than the normal 2n complement of DNA found during the G1 stage of the cell cycle).
Because the characteristic breakage of DNA during apoptosis leads to this accumulation of sub-
Gl events, this measure is often used as an indicator of apoptosis. The results suggested to the
authors that the developmental toxicity of DCA, particularly the induction of embryonic neural-
tube defects, may be mechanistically associated with its ability to increase apoptosis. Bantle et al.
(1999) reported equivocal results for DCA in the Frog Embryo Teratogenesis Assay — Xenopus
(FETAX). Although these in vitro studies cannot be used for dose-response assessment, they
support the in vivo observations that DCA can cause developmental effects.
No single- or two-generation reproductive toxicity studies have been conducted with
DCA. However, there are data indicating that DCA induces male reproductive tract toxicity.
Cicmanec et al. (1991) observed testicular degenerative changes in beagle dogs exposed to 12.5
mg/kg/day DCA in gelatin capsules for 90 days. Testicular changes included syncytial giant cell
formation and degeneration of the testicular germinal epithelium. This dose was the study
LOAEL, based on the testicular changes, liver vacuolization and brain histopathology. Toth et al
(1992) found decreased absolute weight of the preputial gland and epididymis at the lowest dose
tested (31.3 mg/kg/day) in male Long-Evans rats (18 to 19/group) treated with gavage doses of
DCA for 10 weeks, but the absolute testes weight was not affected. Relative liver weights were
also increased in this dose group. At the two higher doses (62.5 and 125 mg/kg/day), significant
decreases in the percentage of motile sperm, epididymal sperm counts and sperm motion
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parameters were observed. Gross pathologic examination did not reveal lesions at any dose in the
testis or epididymal epithelium. Histological evidence of impaired spermiation was noted in the
62.5 and 125 mg/kg/day dose groups and was attributed to retention of late-step spermatids in the
seminiferous tubules. Male fertility was assessed by mating treated males with untreated females
on day 70 (final day of treatment); no effect on fertility was seen at any dose. However, the study
authors indicated that the experimental protocol may have been insensitive for fertility evaluation,
because an overnight mating period was used and multiple matings were possible. Based on
reduced weight in sexual accessory organs (preputial gland and epididymis), the LOAEL for this
study was 31.3 mg/kg/day, and a NOAEL could not be determined.
Linder et al. (1997a) examined the reproductive tract toxicity and spermatotoxicity of
DC A. Male Sprague-Dawley rats (8/group) were given daily gavage doses of 0, 18, 54, 160,
480, or 1440 mg/kg DC A for 14 days. The only evidence of general toxicity was a statistically
significant decrease in final body weight at 480 and 1440 mg/kg/day. A variety of male
reproductive-tract toxicity parameters were affected. Rats in the 54 mg/kg/day group exhibited
clear histopathological effects on spermiation indicative of spermatotoxicity, which increased in
severity with increasing dose. At doses of 160 mg/kg and higher, the percentage of fused cauda
sperm was statistically increased and percent motile sperm were statistically decreased as
compared with controls. The characteristic pattern of response included altered spermiation
(including retention of mature sperm), atypical formation and resorption of residual cytoplasm,
abnormal sperm morphology, and decreased motility, all of which increased in magnitude and
severity with dose. Epididymis weight was decreased at dose levels of 480 and 1440 mg/kg/day.
The LOAEL for this study was 54 mg/kg/day, and 18 mg/kg/day was an equivocal NOAEL. This
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determination is based on unequivocal histopathological changes indicative of spermatotoxicity in
the male reproductive tract at the LOAEL. The susceptibility of the developing male reproductive
tract has not been tested for DCA.
Systemic Effects
Stacpoole et al. (1998) reviewed the clinical data for DCA. The sodium salt of DCA has
been employed experimentally for more than 25 years in clinical medicine and human research.
Over 40 patient-years of cumulative DCA exposure have been recorded in more than 50 infants
and children with congenital lactic acidosis; several patients were treated with daily oral doses of
approximately 25 mg/kg for up to 5 years. Typical doses range from 25 to 100 mg/kg, and
subsequent clinically adverse findings have been mainly restricted to mild neurological effects,
such as a reduction in anxiety or sedative effects in both adults and children that lasted for several
hours. A doses of 50 to 100 mg/kg, there were a few cases of reversible peripheral neuropathy.
No effects on reproductive organs or reproductive parameters have been reported, although the
clinical studies have not specifically targeted evaluation of reproductive toxicity. Limited
pharmacokinetic data in children ranging in age from 18 months to 10 years, who were treated
with DCA for control of lactic acidosis due to malaria, showed kinetic parameters after a single
50 mg/kg infusion similar to those obtained in healthy or acidotic adults (Stacpoole et al., 1998).
Children chronically treated with orally-administered DCA for various congenital forms of lactic
acidosis also exhibited kinetic parameters similar to those of adults undergoing the same treatment
regime (Stacpoole et al., 1998). These human investigations, conducted over more than 25 years
of therapeutic use at pharmacologic doses, suggest that children are no more sensitive to the toxic
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effects of DC A than adults.
As part of a series of neurotoxicity studies, Moser et al. (1999) examined the
neurobehavioral toxicity in weanling and young adults of two rat strains (inbred F344 and outbred
Long-Evans) exposed to DCA in drinking water. Treatment for adult and weanling rats began at
approximately 60 and 21 days of age, respectively. Behavioral and neurological changes were
evaluated using a functional observational battery (FOB) and assessing motor activity. The FOB
consists of a neurobehavioral screening battery comprised of a series of observational and
manipulative tests designed to assess neurological integrity. Male Long-Evans (LE) and F344
weanling rats received 0, 200, 1000, or 2000 mg/L in drinking water (10/dose/strain), and were
tested at weeks 3, 6, 9, and 13. Equivalent daily doses were 0, 17, 88, and 192 mg/kg/day for LE
rats and 0, 16, 89, and 173 mg/kg/day for F344 rats. DCA toxicity was mainly limited to the
neuromuscular system and to endpoints demonstrating altered neuromuscular function; these
included gait abnormalities and changes in grip strength, landing foot splay and righting reaction.
The LOAEL was 16 mg/kg/day for weanling F344 rats and 17 mg/kg/day for weanling LE rats
after 13 weeks of exposure. No NOAELs were identified. Although the LOAELs for the two
strains were similar, the magnitude and severity of neurobehavioral effects were more pronounced
in F344 rats. Thus, F344 weanling rats appear to be more sensitive to DCA-induced
neurotoxicity than LE weanling rats. In the segment of the study using adult rats, the same
experimental design and protocol was used. Male LE and F344 adult rats (10/dose/strain)
received 0, 250, 1250, or 2500 mg/L in drinking water for 8 weeks, and were tested at weeks 2,
5, 8, and 10 (2 weeks after the termination of exposure). Equivalent daily doses were 0, 23, 122,
and 220 mg/kg/day for LE rats and 0, 18, 91, and 167 mg/kg/day for F344 rats. Based on gait
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impairment, the LOAEL for adult F344 rats was 18 mg/kg/day, and there was no NOAEL. For
adult LE rats, the LOAEL was 122 mg/kg/day and the NOAEL was 23 mg/kg/day. Adverse
effects in weanling rats were considered by the authors to be more severe than those in adults
similarly tested and evaluated. The study authors concluded that F344 rats are more sensitive
than LE rats. They also noted that weanlings may be somewhat more sensitive than adults to the
neurobehavioral toxicity of DC A, although this difference was seen only in the LE rats.
The systemic, noncancer effects of DC A in animals and humans can be grouped into four
categories: metabolic alterations, hepatic toxicity, reproductive/developmental toxicity, and
neurotoxicity. The liver is a major target organ of toxicity in DCA-treated mice, rats, dogs and
humans (Cicmanec et al., 1991; DeAngelo et al., 1999; Katz et al., 1981; Parrish et al., 1996;
Stacpoole et al., 1998). Frank, dose-dependent toxicity in the form of hepatic vacuolation,
cytomegaly, karyomegaly, and multi-focal coagulative necrosis has only been observed in mice at
drinking water doses of 500 mg/L ( 84 mg/kg/day) and higher (DeAngelo et al., 1999). Increased
testes weights have been observed in rats dosed with 40.2 mg/kg/day DCA in drinking water for a
lifetime (DeAngelo et al., 1996), although there was no clear dose-response. In a subchronic
study in rats, Katz et al. (1981) reported a LOAEL of 125 mg/kg/day for brain lesions
(vacuolization of the myelinated white tracts); no NOAEL was identified. In a 90-day study in
beagle dogs administered DCA by capsule, Katz et al. (1981) found a LOAEL of 50 mg/kg/day,
based on prostate gland atrophy and testicular changes (degeneration of germinal epithelium,
vacuolation of Ley dig cells, formation of syncytial giant cells). Other reported effects at 50
mg/kg/day and higher included decreased body weight, changes in hematologic and clinical
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chemistry parameters, neurotoxicity, and ocular effects (irreversible lenticular opacities). No
NOAEL could be determined from this study.
In a subchronic study by Cicmanec et al. (1991), four-month old male and female beagle
dogs (5/sex/dose) were administered 0, 12.5, 39.5, or 72 mg/kg/day of dichloroacetate in gelatin
capsules for 90 days. Histopathological testicular changes were reported in the testes of males at
all dose levels (except for controls), and included syncytial giant cell formation and degeneration
of the testicular germinal epithelium. Prostate glandular atrophy characterized by a significant
reduction of glandular alveoli was also noted in mid-and high-dose groups. Other histopathology
included hepatic vacuolization and vacuolization of the myelinated white tracts of the cerebrum
and cerebellum at all dose levels. The LOAEL for this study was 12.5 mg/kg/day, based on
testicular degenerative changes, liver vacuolization and brain histopathology; a NOAEL could
not be determined.
Carcinogenicity
No epidemiological investigations of the carcinogenicity of DC A in humans have been
performed. In animals, there have been a number of independent long-term studies investigating
aspects of the carcinogenicity of DC A. Statistically significant increases in hepatic carcinomas
alone and/or hepatic carcinomas plus adenomas were seen in the following studies: (1) all seven
male B6C3F1 mouse studies (Anna et al., 1994; Bull et al., 1990; DeAngelo et al., 1991;
DeAngelo et al., 1999; Daniel et al., 1992; Ferreira-Gonzalez et al., 1995; Herren-Freund et al.,
1987); (2) two female B6C3F1 mouse studies (Pereira and Phelps, 1996; Pereira, 1996), with the
third female B6C3F1 mouse study reporting hyperplastic nodules in the livers of some treated
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animals, but no liver tumors (Bull et al., 1990); and (3) male F344 rat studies (DeAngelo et al.,
1996; Richmond et al., 1995). The weight of evidence clearly indicates that DCA is
hepatocarcinogenic in both male mice and male rats at high concentrations of DCA given in
water. Exposure levels causing increased incidence of animals with tumors range from 500 to
5000 mg/L. However, concentrations as low as 50 mg/L (8 mg/kg/day) have been reported to
increase the multiplicity of tumors in male mice (DeAngelo et al., 1999). No studies have been
conducted to investigate the transplacental carcinogenicity of DCA or carcinogenicity in young
animals exposed to DCA from birth.
Available data are not adequate to identify any single mode of action as being the only or
most important pathway leading to cancer. DCA has been found to be mutagenic and clastogenic,
but these responses generally occur at high dose levels. There are differences of opinion
regarding genotoxicity of DCA at lower dose levels. IARC (1995) and ILSI (1997) concluded
that it was not genotoxic. DCA was classified as a weak mutagen by Moore and Brock (2000)
and as a direct acting genotoxic agent by the National Center for Environmental
Assessment at EPA (EPA, 1998e) based on recently published studies (DeMarini et al., 1994;
Fuscoe et al., 1996; Leavitt et al., 1997; Harrington-Brock et al., 1998). Possible nongenotoxic
modes of action may contribute to the carcinogenic response but have not been conclusively
implicated in rodent liver tumorigenesis. These MO As include cytotoxicity and reparative cell
proliferation and/or decreased apoptosis, DNA hypomethylation, metabolic alterations involving
glycogen accumulation in hepatocytes, and/or selective growth of subpopulations of mutated or
immunoreactive cells.
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The EPA reviewed DC A carcinogenicity in 1994 (EPA, 1994g), and this evaluation was
updated in 1996 (EPA, 1996d). These reviews classified DC A as a Group B2: Probable Human
Carcinogen, in accordance with the 1986 EPA Guidelines for Carcinogen Risk Assessment. In
1995, IARC concluded that "DCA is not classifiable as to its carcinogenicity to humans," and
placed DCA in the IARC Group 3 category (IARC, 1995). It should be noted that at the time of
the IARC evaluation, there were no data demonstrating high-dose DCA-induced
hepatocarcinogenicity in the rat. In 2002, IARC (2002) changed the cancer classification to
Group 2B (possibly carcinogenic to humans). Under the 1999 Carcinogen Risk Assessment
Guidelines (EPA, 1999), DCA is likely to be carcinogenic in humans, based on the weight of the
evidence in animal bioassays. In particular, the following characteristics were considered and
found to increase the overall weight of evidence for this classification: the number of independent
studies reporting consistently positive results and at roughly comparable doses, site concordance
for tumor formation between two species, consistent observations in different species and sexes,
and clear evidence of a dose-response relationship. There is no clear mechanistic understanding
of the carcinogenic process and the shape of the cancer dose-response curve at low doses. This
precludes any consideration of classifying the tumorigenic dose response for DCA as nonlinear.
The cancer risk from ingestion of DCA was quantified based on a dose-response study in
male mice (DeAngelo et al., 1999). The cumulative incidence of hepatic total tumor incidence
(carcinoma plus adenoma) in the test animals was well-described by several different dichotomous
models, with the multistage model yielding the best fit. Based on this model, the BMD10 was 6.86
mg/kg/day, and the BMDL10 was 2.1 mg/kg/day. In accord with the draft guidelines (U.S. EPA,
1999), the BMDL10 was used as the point-of-departure (POD) for quantifying cancer risk.
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Because the mode of action by which DCA increases cancer risk is not understood, extrapolation
to low dose was performed by assuming a no-threshold linear dose-response curve between the
origin and the POD. This yields a cancer slope factor of 0.048 per mg/kg/day and a unit risk of
1.4 x 10"6 per (|ig/L). Based on this calculation, drinking water concentrations of approximately
70 |ig/L, 7 |ig/L and 0.7 |ig/L are estimated to be associated with lifetime cancer risks of 10"4, 10"5
and 10"6, respectively.
Basis for RfD and MCLG
EPA based its RfD on the LOAEL of 12.5 mg/kg/day identified for testicular degenerative
changes, liver vacuolization and brain histopathology in the Cicmanec et al. (1991) study. EPA
used an UF of 3000, for use of a LOAEL, inter-and intraspecies variability, and use of a study
with a less-than-lifetime duration (EPA, 2001c). An RfD of 0.004 mg/kg/day results,
corresponding to a DWEL of 0.14 mg/L (rounded to 0.1 mg/L) for a 70 kg adult drinking 2 L of
water per day.
The MCLG for DCA is zero, based on its probable carcinogenicity and the use of linear
low-dose extrapolation.
Children's Risk in Relation to the MCLG
The Agency believes that the MCLG of zero is protective for both the carcinogenic and
systemic effects of DCA in children and adults. In addition, there is no evidence from studies on
the systemic effects of DCA that children are more sensitive to the toxic effects of DCA than are
adults. Human investigations, conducted over more than 25 years of therapeutic use at
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pharmacologic doses, suggest that DCA toxicokinetics do not differ between children and adults,
and that children are not likely to be more vulnerable to the potentially toxic effects of DCA
(Stacpoole et al., 1998). However, these data are limited and no definitive conclusions can be
made.
No multi-generation reproductive toxicity animal studies have been conducted to assess
age-related differences in sensitivity to DCA. However, in female rats, DCA exposure during
gestation resulted in the impairment of fetal maturation and soft tissue anomalies (primarily of
cardiac origin) indicating that the developing fetus is susceptible to DCA-induced toxicity. Data
collected by Moser et al. (1999) provide limited evidence for increased susceptibility of rats to
DCA-induced neurotoxicity when exposures begin shortly after weaning.
Children with certain inborn errors of metabolism may be at increased risk for DCA-
induced toxicity. Glycogen-storage disease is an inherited deficiency or alteration in any one of
the enzymes involved in glycogen degradation. The liver is a major target organ of toxicity in
DCA-treated mice, rats, dogs, and humans, and dose-related increases in liver size have been
reported to be accompanied by an increase in glycogen deposition in the liver in mice and rats
(Kato-Weinstein et al., 1998). Although the enzymatic basis and functional significance of this
finding is unclear, glycogen accumulation may be associated with inhibition of glycogenolysis, as
the reported effects resemble those observed in glycogen-storage disease VI (EPA, 2001c). If
glycogen accumulation plays a role in the DCA tumorigenic process, children with this disease
may represent a group that is more sensitive to DCA toxicity. Hereditary tyrosinemia (a human
disease involving a deficit in tyrosine metabolism) is often associated with the development of
hepatocellular carcinoma in young patients (Tanguay et al., 1996; LaBerge, 1986). DCA has
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been reported to significantly alter tyrosine metabolism as a consequence of its inhibitory effect on
glutathione S-transferase Zeta (Cornett et al., 1999). DCA-induced inhibition of tyrosine
metabolism may result in increased levels of reactive tyrosine metabolites that may adversely
affect the liver and nervous system. Children with hereditary tyrosinemia may be at increased risk
for DCA-induced toxicity. Oxalate is an excretory endpoint of DC A metabolism. Thus, children
affected with the metabolic disorder, hyperoxaluria, may also be at an increased risk to DCA
exposure. Adults with hereditary tyrosinemia, glycogen storage disease, or hyperoxaluria may
also be at increased risk for DCA-induced toxicity.
2.3.3. Trichloroacetic Acid
Developmental/Reproductive Effects
Johnson et al. (1998) studied the teratogenicity of TCA by exposing pregnant Sprague-
Dawley rats to 0 or 2,730 mg/L TCA in neutralized drinking water on gestation days 1-22. The
authors estimated the doses to be equivalent to 0 or 291 mg/kg/day. Dams were sacrificed on
gestation day 22; implantation sites, resorption sites, fetal placements, fetal weights, placental
weights, crown-rump lengths, gross fetal abnormalities, and abnormal abdominal organs were
recorded. In addition, the fetal hearts were removed, dissected, and examined for abnormalities
under microscope. There was no microscopic evaluation of other internal anomalies, and no
evaluation of skeletal anomalies. Although the authors reported no signs of maternal toxicity and
no effect on maternal weight gain, the average maternal weight gain for TCA-exposed animals
was 84.6 g as compared with 122 g for control animals, representing a 30% decrease in maternal
body weight gain. This decrease is considered toxicologically significant. Significant increases
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were observed in average resorption sites and average implantation sites. No significant
differences were found in the numbers of live or dead fetuses, fetal weight, placental weight,
crown-rump length, external morphology, or gross external or noncardiac internal congenital
abnormalities. Cardiac abnormalities were evident in 10.5% of the fetuses in the TCA group,
compared to 2.15% of the controls. No single cardiac defect or group of defects predominated in
the TCA group. Based on the increases in implantation and resorption sites, and cardiac
malformations in the single TCA-treated group, 291 mg/kg/day is considered a developmental
LOAEL. Based on decreased maternal weight gain, the single dose of 291 mg/kg/day is also a
maternal LOAEL.
In a study by Smith et al. (1989), pregnant Long-Evans rats (20 animals/dose) received
TCA at doses of 0, 330, 800, 1200, or 1800 mg/kg/day in drinking water during gestation days 6-
15. Maternal spleen and kidney weights were increased significantly in all dose groups in a dose-
dependent manner (p=0.0001); liver weights of dams were not affected by TCA treatment.
Postimplantation loss increased at doses of 330 mg/kg/day and higher. Fetal body weight and
crown-rump length were significantly (p<0.05) lower than controls for all dose groups. Soft-
tissue malformations in the cardiovascular system were increased for all treatment groups in a
dose-dependent manner. Levocardia occurred in 0%, 32%, 71%, 71%, and 88% of the litters in
the 0, 330, 800, 1200, and 1800 mg/kg/day groups, respectively. The lowest dose, 330
mg/kg/day, was considered the LOAEL in this study, based on the dose-dependent maternal
effects (increased kidney and spleen weights) and developmental effects (decreased fetal weight
and crown-rump length and increased incidences of levocardia in litters).
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TCA has also been tested in a number of alternative screening models for assessing
potential developmental toxicity. Saillenfait et al. (1995) exposed explanted embryos from
Sprague-Dawley rats on gestational day 10 for 46 hours to TCA concentrations up to 6.0 mM.
TCA induced statistically significant, concentration-related decreases in the growth and
development parameters of conceptuses. Hunter et al. (1996) conducted a 24-hour exposure of
CD-I mice embryos to TCA concentrations up to 5000 |iM. TCA produced abnormal embryonic
development at concentrations that were not embryolethal. In addition, Fort et al. (1993)
reported that TCA induced malformations at doses lower than those that induced lethality in the
FETAX assay.
Conflicting results were reported by Fu et al. (1990), who studied the developmental
toxicity potential of TCA using a regeneration assay from reaggregated Hydra cells. Based on
similarity in the minimal effective toxic concentration for adults and artificial embryos, TCA was
not considered to interfere with development. As previously noted, the Hydra system is designed
to overestimate developmental hazard potential and is considered to be more sensitive to
developmental toxicity than most in vitro mammalian test systems;
Overall, the in vivo studies indicate that TCA can cause developmental toxicity at
maternally toxic doses. The in vitro studies cannot be used quantitatively, but the results of
Saillenfait et al. (1995) and Hunter et al. (1996) support the findings of the in vivo studies. In
contrast, the findings of Fu et al. (1990) suggest that TCA would not be considered a priority
compound for further developmental toxicity testing in vivo.
One in vitro study suggested that TCA might decrease fertilization. The effect of TCA on
in vitro fertilization was examined in hybrid C57BL6 x DBA/2 (B6D2F1 ) mice (Cosby and
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Dukelow, 1992). TCA was constituted in culture medium to yield concentrations of 100, 250, or
1000 ppm (mg/L), and incubated with mouse oocytes and sperm for 24 hours. The percent of
oocytes fertilized was significantly decreased at 250 mg/L (p<0.025) and 1000 mg/L (p<0.1)
compared to controls.
Systemic Effects
Toxicity studies of orally-administered TCA have primarily identified the liver and kidney
as target organs. Hepatic peroxisome proliferation has been examined, with mice reported to be
more sensitive to this effect than rats. Parrish et al. (1996) reported that TCA induced
peroxisome proliferation in B6C3F1 mice exposed for 10 weeks to doses as low as 25
mg/kg/day. Increased liver weight and significant increases in hepatic labeling (a measure of cell
division) have been observed in short-term studies in mice at doses as low as 100 mg/kg/day
(Dees and Travis, 1994). In a 90-day study by Mather et al. (1990) in Sprague-Dawley rats, the
NOAEL was 36.5 mg/kg/day, and the LOAEL was 355 mg/kg/day for reduced spleen weight,
increased relative kidney weight, and increased liver weight that was accompanied by increased
hepatic peroxisomal beta-oxidation activity. Bull et al. (1990) treated groups of mice with TCA
in their drinking water at 0 or 1000 mg/L for 52 weeks, at 2000 mg/L for 37 weeks with a 15-
week recovery period, or at 2000 mg/L for 52 weeks. Small increases in liver size, some
accumulation of lipofuscin and focal necrosis were seen in all groups. The LOAEL for hepatic
lesions from this study was 164 mg/kg/day. In rats exposed to TCA for up to 104 weeks
(DeAngelo et al., 1997), the NOAEL for liver toxicity was 32.5 mg/kg/day and the LOAEL was
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364 mg/kg/day. Liver effects included increased serum levels of liver enzymes (indicating leakage
from cells) and histopathological evidence of necrosis.
Carcinogenicity
TCA induces liver tumors in mice but not in rats. Pereira (1996) observed an increased
incidence of hepatic adenomas and carcinomas in female B6C3F1 mice at doses of 262
mg/kg/day and higher in drinking water for 82 weeks. In contrast, no increase in neoplastic liver
lesions were found in F344 rats given doses up to 364 mg/kg/day for 104 weeks (DeAngelo et al.,
1997). An earlier study by DeAngelo and colleagues (1991) in B6C3F1 mice found hyperplastic
nodules and hepatocellular tumors, both adenomas and carcinomas, primarily in males. The
authors noted that the female mice appear to be less sensitive than the male mice to the
carcinogenic potential of TCA. Bull et al. (1990) found that exposure to TCA via drinking water
resulted in induction of liver tumors in male B6C3F1 mice; female mice, however, did not show
these effects after 52 weeks of exposure.
The mechanism for TCA-induced mouse hepatocarcinogenesis has not been conclusively
determined. Mutagenicity data have provided mixed results (DeMarini et al., 1994; Giller et al.,
1997; Harrington-Brock et al., 1998; reviewed in EPA, 1994g). Evidence for DNA stand breaks
and clastogenicity is also mixed (Nelson and Bull, 1988; Chang et al., 1991). A recent study
found that chromosome damage is not induced by TCA in the absence of pH changes (Mackay et
al., 1995), but Harrington-Brock et al. (1998) found weakly positive evidence of TCA
clastogenicity (small colonies) in mouse lymphoma cells in the absence of pH changes.
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A variety of recent mechanistic studies have observed that TCA induced or promoted liver
tumors in mice (Ferreira-Gonzalez et al., 1995; Latendresse and Pereira, 1997; Pereira and
Phelps, 1996; Stauber and Bull, 1997; Tao et al., 1996, 1998). In a recent review, Moore and
Harrington-Brock (2000) evaluated the weight-of-evidence for the genotoxicity of
trichloroethylene and its metabolites, including TCA. The authors concluded that it is unlikely
that TCA contributes to tumor formation through a mutational mechanism. Moreover, laboratory
mice have a high background rate of spontaneous liver tumor formation, and chemical compounds
are thought to increase this incidence via a nongenotoxic mode of action (Bull, 2000). A variety
of recent mechanistic studies have observed that TCA induced or promoted liver tumors in mice
(Ferreira-Gonzalez et al., 1995; Latendresse and Pereira, 1997; Pereira and Phelps, 1996; Stauber
and Bull, 1997; Tao et al., 1996, 1998). However, laboratory mice have a high background rate
of spontaneous liver tumor formation, and chemical compounds that increase this incidence are
thought to exert their effects via nongenotoxic mode(s) of action (Bull, 2000).
A variety of other mechanisms have been suggested as contributing to TCA-induced liver
tumorigenesis. Of these, peroxisome proliferation and altered regulation of cell growth have been
most well supported. There is little evidence for a role of oxidative DNA damage (Parrish et
al.,1996), or regenerative hyperplasia (Pereira, 1996; DeAngelo et al., 1997; Bull, 2000).
Peroxisome proliferation is activated in both mice and rats, but liver tumors are only induced in
mice (EPA, 1994; Pereira, 1996; DeAngelo et al., 1997; Bull, 2000). The lack of tumorigenicity
in F344 rats (DeAngelo et al., 1997) might reflect a lower affinity of the peroxisome proliferation
pathway for TCA, which would result in a smaller peroxisome response in rats as compared to
mice. Humans have been reported to have a much lower response to exposure to peroxisomal
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proliferators, including TCA, than either mice or rats (Lapinskas and Corton, 1999; Bentley et al.,
1993; Walgren et al. 2000). However, it is not yet clear whether peroxisome proliferation is a key
event in the development of TCA-induced mouse hepatocarcinogenesis. A better case can be
made for altered proliferation in subpopulations of cells having selective growth advantages
(Stauber et al., 1998), arising for example, due to spontaneous mutations (Ferreira-Gonzalez et
al.,1995).
Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, TCA is best
classified as Group C: Possible Human Carcinogen. In accordance with EPA's Draft 1999
Guidelines for Carcinogen Risk Assessment, the data on TCA provide suggestive evidence of
carcinogenicity, but the data are not sufficient to assess human carcinogenic potential. These
conclusions are based on the lack of genotoxicity of TCA in numerous studies, the induction of
only liver tumors in one rodent species (mouse), the uncertainty regarding the likely mode(s) of
action of TCA-induced hepatocarcinogenicity, and the questionable human relevance of the
finding of increased liver tumors in a rodent species with a high background rate of
spontaneously-occurring liver tumors. Based on this same line of reasoning, the data are
insufficient to conduct a dose-response quantification for cancer.
Basis for the RfD and MCLG
EPA based the RfD on the NOAEL of 32.5 mg/kg/day for liver effects in the DeAngelo et
al. (1997) study. A composite UF of 1000 was applied to the NOAEL, based on a 10-fold factor
for extrapolation from an animal study to humans, a 10-fold factor to account for variation in
sensitivity among members of the human population, and a 10-fold factor for data base
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insufficiencies, including lack of adequate developmental toxicity studies in two species, lack of a
multi-generation reproductive study, and lack of full histopathological data in a second species.
The resulting RfD is 0.03 mg/kg/day, corresponding to a DWEL of 1.05 mg/L for a 70 kg adult
drinking 2 L of water per day.
Assuming a drinking water RSC of 20% of total exposure, the MCLG = 1.05 mg/L x 0.2 /
10 to account for possible carcinogenicity = 0.021 mg/L, rounded to 0.02 mg/L. The default
RSC of 20% was chosen in accordance with the exposure decision tree approach in EPA's
Human Health Methodology (EPA, 2000f), taking into account the likelihood of exposure to
TCA from sources other than tap water, such as ambient air and food. The available data are
sufficient to demonstrate that food and air are relevant exposure sources in addition to drinking
water, but the data are inadequate to calculate an RSC.
Children's Risk in Relation to the MCLG
The MCLG is expected to be protective of fetuses and children. Developmental and
systemic toxicity appear to occur at similar doses (i.e. similar LOAELs), although conclusions are
limited by the lack of developmental NOAELs. For example, LOAELs for developmental toxicity
have been reported as 291 mg/kg/day (Johnson et al, 1998) and 330 mg/kg/day (Smith et al.,
1989). The developmental toxicity LOAELs are comparable to the LOAEL of 364 mg/kg/day for
systemic toxicity in the study used to derive the RfD (DeAngelo et al., 1997). The MCLG based
on systemic effects is also likely to protect children because remaining uncertainties regarding the
NOAEL for developmental toxicity are taken into account through the application of UFs. In
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addition, there is no evidence from studies on the systemic effects of TCA that children are more
sensitive to the toxic effects of TCA than are adults.
2.3.4. Monobromoacetic Acid
Developmental/Reproductive Effects
No peer-reviewed developmental toxicity studies of MBA are available. In a published
abstract, Randall et al. (1991) reported on the reproductive and developmental toxicity of MBA.
Pregnant Long-Evans rats were given oral gavage doses of 0, 25, 50, or 100 mg/kg/day MBA in
distilled water on gestation days 6-15. In the high-dose group, maternal weight gain was reduced
and one dam died. No effects on reproduction were observed. Several developmental effects
were noted in the high-dose group, including decreased size of live fetuses (the affected measure
of size was not provided in the study summary) and increased incidence of soft tissue
malformations, most of which were cardiovascular and craniofacial. Based on the limited data
provided in the abstract, the NOAEL for this study was 50 mg/kg/day and the LOAEL for both
maternal and developmental effects was 100 mg/kg/day.
The potential developmental toxicity of MBA has been evaluated in whole embryo culture,
and malformations were increased at sub-lethal doses (Hunter et al., 1996). These whole-embryo
testing data provide support for the developmental toxicity of the brominated acetic acids
observed in vivo.
Linder et al. (1994a) reported the results of acute toxicity and acute spermatotoxicity
studies of MBA. In the spermatotoxicity study, groups of eight male Sprague-Dawley rats were
given single doses of 0 or 100 mg/kg MBA in a volume of 5 mL/kg and were sacrificed 2 or 14
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days after dosing. The selected single dose of 100 mg/kg was an approximate LD01, and was
chosen to provide a relatively high dose with a minimal likelihood of mortality. Measures of male
reproductive toxicity included reproductive organ weights, sperm counts, sperm morphology,
sperm motility, and histopathological examination of the seminiferous tubules. No adverse effects
were observed in the single-dose study; therefore, a repeated-dosing protocol experiment was also
conducted. Groups of eight rats were given daily doses of 0 or 25 mg/kg/day MBA in water for
14 days and were sacrificed 24 hours after the last dose. MBA did not induce any
spermatotoxicity in this repeated-dosing study.
Systemic Effects
The toxicity data for MBA are very limited. The oral LD50 for MBA was reported as 177
mg/kg in male rats (Linder et al., 1994a). Single-dose (0 or 100 mg/kg) and 14-day studies (0 or
25 mg/kg/day) have been conducted to assess the spermatotoxicity of MBA. No general toxicity
was observed with either dosing regimen (Linder et al., 1994a). No data were identified on young
animals to compare the potential susceptibility of children and adults to the toxic effects of MBA.
Carcinogenicity
No data were identified on the carcinogenicity of MBA. The genotoxicity data for MBA
have provided mixed results. MBA was mutagenic in S. typhimurium (Giller et al., 1997; Kohan
et al., 1998; NTP, 2000a) and induced DNA single-strand breaks in vitro (Stratton et al., 1981),
but did not induce a DNA repair system (SOS DNA repair) in Escherichia coli that responds to
primary DNA damage, and did not cause micronuclei in a newt larvae system (Giller et al., 1997).
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Based on the absence of human or animal data and only equivocal genotoxicity data, MBA is
classified as Group D: Not Classifiable as to Human Carcinogenicity under the 1986 Cancer Risk
Assessment Guidelines. Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment
(EPA, 1999), the data on MBA are inadequate for an assessment of human carcinogenic
potential.
Basis for RfD and MCLG
The data on MBA are inadequate to derive an RfD, because there are no systemic toxicity
studies of sufficient duration. Similarly, there are insufficient data to assess the carcinogenic
potential of MBA. In the absence of adequate data on the noncancer or cancer effects of MBA,
no MCLG is proposed.
Children's Risk in Relation to the MCLG
Data relevant to potential fetal sensitivity are limited to a single developmental study
reported in a published abstract (Randall et al., 1991). In rats administered MBA on GD 6-15,
the NOAEL was 50 mg/kg/day for both maternal effects (decreased maternal weight gain) and for
fetal effects (decreased live-fetus size and increased incidence of soft-tissue malformations). The
induction of developmental effects only at doses that also affected maternal body weight does not
suggest that the fetus is more sensitive to the toxic effects of MBA.
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2.3.5. Bromochloroacetic Acid
Developmental/Reproductive Effects
No standard peer-reviewed developmental toxicity studies of BCA are available. NTP
(1998b) reported the results of a short-term reproductive and developmental toxicity screening
protocol for BCA. Two groups of male rats and three groups of female rats were treated with
BCA at concentrations of 0, 60, 200, or 600 mg/L. In males, exposed either on study days 6-35
or exposed on study days 6-31 followed by a BrdU treatment, no consistent treatment-related
effects on epididymal sperm measures, spermatid head counts, sperm morphology or sperm
motility were observed in either group at necropsy. Group A females were treated prior to
mating, during mating to treated males, and during early gestation. Group B females were mated
with treated males, and exposed on gestation days. Group C females were treated with BCA
prior to mating and were subsequently treated with BrdU. No effects on indices of mating or
fertility were affected in Group A or C females. Analysis of the results after combining data for
groups A and C revealed statistically significant decreases to 70% of controls for live fetuses per
litter and to 75% of controls in total implants per litter at 50 mg/kg/day. For group B females,
increased (but not statistically significant) post-implantation losses, and total resorptions were
observed. No treatment-related effects were observed upon soft tissue examination (heart and
brain) of the fetuses. The NOAEL for statistically significant decreases in fertility was the mid
dose, 19 mg/kg/day (200 ppm treatment group). The LOAEL for reproductive and
developmental effects (decreased implants per litter and live fetuses per litter) was the high dose
of 50 mg/kg/day (600 ppm treatment group).
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In a study submitted for publication, Klinefelter et al. (2002) administered BCA (dissolved
in deionized water and pH-adjusted) by gavage to adult male Sprague-Dawley rats (12/dose) at
doses of 0, 24, 72, or 216 mg/kg/day for 14 days. Body weight was significantly decreased in the
highest dose group. Testis, epididymis, and seminal vesicle weights were unaffected by BCA
treatment, and there was no effect on testis sperm production or serum testosterone. Although
spermatid numbers were not altered by BCA exposure, a significant dose-related decline in
epididymal sperm reserves was observed at 72 and 216 mg/kg/day, with the effect on cauda
epididymal sperm being more severe than on caput epididymal sperm. Dose-related decreases in
serum luteinizing hormone (LH), follicle-stimulating hormone (FSH), and prolactin were noted in
all dosed groups, with statistical significance occurring in the two highest dose groups. Dose-
related decreases were also observed in the percentage of motile and progressively motile cauda
sperm, and in sperm motion parameters (i.e., velocity and linearity) and the percentage of
morphologically normal cauda and caput sperm. Caput sperm abnormalities were characterized
by an increased number of sperm with misshapen heads or tail defects, whereas cauda sperm
abnormalities consisted mainly of an increased number of isolated heads. Histological evaluation
of the testis showed a dose-related increase (statistically significant in the two highest dose
groups) in the number of Step 19 spermatids retained in Stage X and XI of the spermatogenic
cycle. Other findings included a dose-related increase in the number and size of atypical residual
bodies in Stages X and XI (not quantified) and a shift in localization of these bodies, from basal
migration to luminal release, with increasing BCA dose. According to the study authors, the
LOAEL for altered spermiation in this study was 24 mg/kg/day, and a NOAEL could not be
determined.
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In a subsequent study in the same report (Klinefelter et al., 2002), adult male Sprague-
Dawley rats (10/dose) were administered 14 daily gavage doses of BCA (dissolved in deionized
water and pH-adjusted) of 0, 8, 24, or 72 mg/kg/day. End points evaluated were the same as
those assessed in the previous study. Additionally, sperm protein was extracted and analyzed, and
a fertility assessment was conducted via in utero insemination of untreated females (synchronized
by administering a subcutaneous injection of an agonist for luteinizing hormone releasing
hormone, LHRH) with sperm from treated males. Inseminated females were sacrificed 9 days
following treatment, and implanted fetuses and corpora lutea of pregnancy were counted. No
treatment-related changes in body weight, testes weight, or the weight of the seminal vesicles
were observed. However, in contrast with the previous study conducted by the same authors,
epididymal weights were reduced at 72 mg/kg/day, and there were no differences between treated
and control groups in any of the hormonal measurements. Sperm motion parameters were
consistently altered by BCA exposure. Although the percentage of motile sperm was only
decreased in the high-dose group (72 mg/kg/day), progressive sperm motility was decreased at all
doses tested. Altered sperm morphology was only observed at 72 mg/kg/day; abnormalities in
both cauda and caput sperm were similar to those observed in the earlier study. The cauda sperm
showed increased incidences of tail defects and the caput sperm showed increased incidences of
isolated heads. In utero insemination of untreated females with the cauda epididymal sperm from
treated males showed a significant reduction in fertility at all doses, but there was no dose-
response. Fertility rates in the 8, 24, and 72 mg/kg/day groups were 33%, 44%, and 37%,
respectively, as compared with 75% in control animals. Decreased levels of the sperm protein
SP22 correlated with the reduction in fertility, supporting the conclusion that SP22 is a useful
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sperm biomarker of fertility. The LOAEL for this study was 8 mg/kg/day, the lowest dose tested,
and a NOAEL could not be determined.
The effects of BCA on male reproduction have also been evaluated following oral gavage
dosing in mice. Luft et al. (2000) reported in an abstract on a study in which male C57BL/6 mice
(12 mice/group) were administered daily gavage doses of 0, 8, 24, 72, or 216 mg/kg BCA for 14
days. After 14 days, five mice/group were necropsied for histopathological examination of the
testes, epididymis, and seminal vesicles. The remaining seven males were used in a 40-day
breeding assay to evaluate the effects of treatment on fertility. Coital plug-positive females
(presumably untreated) were replaced daily, and uteri were dissected 14 days later; the numbers of
implantations, resorptions and fetuses were determined. No effects on body weight or
reproductive organ weights were observed for any of the dose groups. Results of
histopathological examination of the male reproductive tissues were not reported. BCA
treatments with 72 or 216 mg/kg/day resulted in adverse reproductive performance, but only for
the first ten days after treatment (data not shown). Significantly decreased measures of
reproductive performance included: mean number of litters per male; percentage of litters per
female bred, as measured by the percent of plug-positive females that became pregnant; and total
number of fetuses per male. There was no difference in the number of coital plugs, suggesting
that treatment did not result in behavioral effects on mating. The number of fetuses per litter,
number of resorptions, and number of terata were not affected, suggesting that, under the
conditions of this study, adverse reproductive effects on the male did not induce developmental
toxicity. This study appears to have identified a NOAEL of 24 mg/kg/day and a LOAEL for
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decreased male fertility of 72 mg/kg/day, but a definitive conclusion is not possible until the full
study (rather than just an abstract) is published.
Andrews et al. (1999), in a published abstract, treated gestational day 9 embryos to BCA
concentrations up to 300 |iM BCA for 48 hours and then scored for dysmorphology,
developmental score, head length, somite number, crown-rump length and embryo lethality.
Treatment with BCA at 300 |iM or greater was dysmorphogenic; and BCA at 200 |iM or greater
significantly affected the developmental score, head length and somite number. In another whole
embryo culture study designed to test mechanisms of haloacetic acid-induced dysmorphogenesis,
Ward et al. (2000) reported that BCA treatment increased the accumulation of sub-Gl events (a
measure of cells with less than the normal 2n complement of DNA found during the G1 stage of
the cell cycle). Because the characteristic breakage of DNA during apoptosis leads to this
accumulation of sub-Gl events, this measure is often used as an indicator of apoptosis. The
apoptotic response was also confirmed by fluorescence microscopy using an acidophilic dye. The
ability of BCA to induce an apoptotic response suggested to the authors that the developmental
toxicity of BCA, particularly the induction of embryonic neural-tube defects, is mechanistically
associated with its ability to increase apoptosis. The whole embryo testing data provide
mechanistic support for the developmental toxicity of the brominated acetic acids observed in
vivo.
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Systemic Effects
Oral toxicity studies have identified the kidney and liver as potential target organs for
BCA toxicity. Parrish et al. (1996) administered male mice 0, 25, 125, or 500 mg/kg/day BCA in
drinking water for 21 days. Administered doses were estimated from drinking water
concentrations based on default water intake values for male B6C3F1 mice (EPA, 1988).
Increased liver weight was induced by the highest dose tested (500 mg/kg/day), with a NOAEL of
125 mg/kg/day. NTP (1998b) evaluated target organ toxicity as part of a reproductive and
developmental screening assay in rats. The NOAEL was 15 mg/kg/day, and the LOAEL was 39
mg/kg/day for treatment-related liver histopathological changes (cytoplasmic vacuolization) and
increased liver weight. There is no evidence that children are more sensitive to the toxic effects of
BCA than are adults. However, these data are limited and no definitive conclusions can be made.
Carcinogenicity
In a published abstract, Stauber et al. (1995) reported that BCA induces liver tumors in
B6C3F1 mice. There are no published reports of a full bioassay with BCA. BCA was mutagenic
in S. typhimurium (NTP, 2000b) and induced oxidative DNA damage in the livers of mice given
treated drinking water (Parrish et al., 1996). Based on the absence of human or animal data and
very limited genotoxicity data, BCA is classified as Group D: Not Classifiable as to Human
Carcinogenicity under the 1986 Carcinogen Risk Assessment Guidelines. Under EPA's Draft
1999 Guidelines for Carcinogen Risk Assessment (EPA, 1999), the data on BCA are inadequate
for an assessment of human carcinogenic potential.
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Basis for RfD and MCLG
The data on BCA were insufficient for the derivation of an RfD, because there are no
systemic toxicity studies of sufficient duration. Similarly, there are insufficient data to assess the
carcinogenic potential of BCA. In the absence of adequate data on the noncancer or cancer
effects of BCA, no MCLG is proposed.
Children's Risk in Relation to the MCLG
Data on the effects of BCA on fetuses and children are very limited. NTP (1998b)
conducted a reproductive and developmental toxicity screening assay. The NOAEL for decreased
live fetuses/litter and decreased total implants/litter was 19 mg/kg/day, while similar doses were
considered the NOAEL for general toxicity in both adult males (15 mg/kg/day) and females (19
mg/kg/day). Decreased fertility was observed when adult males were exposed to doses as low as
8 mg/kg/day; no NOAEL was identified (Klinefelter et al., 2002). Although this latter endpoint is
a reproductive effect, the LOAEL is based on exposure of the adult male; no data were identified
on whether exposure of young males enhances their sensitivity. Thus, the data are limited, but do
not support the hypothesis that fetuses or children are more sensitive than adults.
2.3.6. Dibromoacetic Acid
Developmental/Reproductive Effects
Two recent studies have evaluated the reproductive and developmental toxicity of DBA in
Sprague-Dawley rats. Christian et al. (2001b) administered DBA in deionized drinking water to
male and female rats (10/sex/group) at concentrations of 0, 125, 250, 500, or 1000 ppm,
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beginning 14 days prior to cohabitation and continuing through gestation and lactation (63-70
days of treatment). The average daily doses (based on measured water consumption and body
weights) varied, depending on the phase of reproduction. Among the pups, two male and two
female weanlings from each litter were selected for one additional week of observation
(postweanling days 1-8, commencing on LD 29); daily food intake, drinking water consumption,
and body weights were recorded, and necropsy was conducted at sacrifice. Apparent taste
aversion was associated with an exposure-dependent reduction in water consumption, which was
paralleled by a reduction in food intake at all concentrations. Decreased body weight gain was
observed in parental animals and postweanling pups at the two highest exposure levels. Estrous
cycling was unaffected in the female rats. The only observed adverse reproductive effect was a
possible reduction in mating performance in the 1000 ppm group, as evidenced by a slight but
nonsignificant increase in the number of days of cohabitation and a decrease in the number of
mated pairs (6/10 in the 1000 ppm group versus 9-10/10 in all other groups). There were no
effects on pre- and postimplantation losses, live litter sizes, and gross external morphology or sex
ratios in the pups. Although an exposure-related decrease in pup body weights was noted, these
findings were attributed to decreased water and food consumption resulting from the poor
palatability of DBA-treated drinking water. Based on a lack of statistically significant, treatment-
related findings, the parental and reproductive/developmental NOAEL for this study is 1000 ppm
(the highest dose tested), and a LOAEL could not be determined. Based on measured water
intake and body weights, the NOAEL corresponds to a dose of 66 mg/kg/day in parental males,
>60 mg/kg/day in parental females, and >82 mg/kg/day for developmental effects.
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The Chlorine Chemistry Council (CCC, 2001) recently completed a two-generation
drinking water study of DBA in rats. The report has been peer reviewed by an EPA scientific
advisory group that has evaluated and accepted the results. Male and female Crl:CD Sprague-
Dawley rats (30/sex/exposure group) were administered DBA in drinking water at concentrations
of 0, 50, 250, or 650 ppm continuously from initiation of exposure of the parental (P) generation
male and female rats through weaning of the F2 offspring. The high concentration was chosen
based on a range-finding study that found that 650 ppm was the highest concentration expected to
allow survival of the Fx offspring. For the P generation, DBA exposure was initiated at 43 days
of age and continued from premating until study day (SD) 92 for males, and from premating
through gestation and a 29-day period of lactation (LD 1-29) for females. Parental generation
offspring (F, males and females) were exposed in utero during gestation, and during lactation
(LD 1-29); selected Fx males and females (30/sex/exposure group) were further exposed during a
postweaning period of at least 71 days, which continued through mating, gestation, and lactation.
All other F, pups were sacrificed on LD 29. All F, adult females and their offspring (F2
generation) were sacrificed on LD 29.
Water consumption was statistically significantly decreased at all exposure levels,
presumably due to taste aversion, and food intake was significantly reduced at the two highest
exposure concentrations. Body weights and body weight gains for high-dose P males and females
were significantly reduced during the premating period and were significantly decreased for high-
dose P females during gestation and lactation. F, male and female pups had significantly reduced
body weights at all exposure levels, sufficient for the study authors to delay weaning until LD 29
to ensure pup survival. By LD 29, the body weights of pups in the 50 ppm group were similar to
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control pup weights. Throughout the postweaning/premating period, Fx males and females in the
250 and 650 ppm groups weighed significantly less than controls, and the females continued to
exhibit significant reductions in body weight (compared to controls) during gestation and
lactation. The body weights of F2 pups in the two highest dose groups were also reduced;
however, these reductions were not reported to be statistically significant, relative to control
group values, by the end of lactation. No treatment-related effects were reported in either
generation for estrous cycling, number of days in cohabitation, duration of gestation, fertility
index, gestation index, number and sex of offspring per litter, number of implantation sites, litter
size, lactation index, percent pup survival, litter sex ratio, and gross malformations. In the 650
ppm group, preputial separation was significantly delayed in the Fx male rats (50.5 days versus
48.1 days in controls), and vaginal patency in F, female rats was also significantly retarded (36.3
days versus 33.4 days in controls); no significant differences were seen, however, when the data
were analyzed using body weight as a covariate. These effects were considered to be due to a
general retardation of growth associated with the significant reduction in body weight in this
exposure group at weaning. In F2 male and female pups, anogenital distance did not differ from
controls on LD 1 but was significantly reduced in male pups in the 250 and 650 ppm by LD 22;
these findings were also considered to be associated with a general retardation of growth rather
than being treatment-related.
An increased incidence of malformations of the male reproductive tract, including small
testes and small or absent epididymides, was observed in four males in the F, group exposed to
650 ppm and was considered to be treatment-related. Histopathologic examination of
reproductive organs of P and F, male rats in the 250 and 650 ppm groups (N =
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30/group/generation) showed a consistent and significant exposure-related increase in retained
Step 19 spermatids in Stage IX and X tubules and in increased or abnormal residual bodies in
affected seminiferous tubules. Diffuse testicular atrophy and phagocytized Step 19 nuclei in the
basilar area of affected seminiferous tubules were also observed, although at a lower incidence.
Other testicular abnormalities in 250 and 650 ppm male rats of both generations included
increased amounts of exfoliated spermatogenic cells/residual bodies in epididymal tubules,
atrophy, and hypospermia. No effects were found on percent motile sperm, sperm count, sperm
density, and number and percent of morphologically abnormal sperm for exposed groups.
An increase in the incidence and intensity of extramedullar hematopoiesis in the red pulp
of the spleen occurred in the Fx generation female rats in the 650 ppm group and may have been
treatment-related. Decreased cellularity of the cortical lymphoid area of the thymus was noted in
P generation females in the two highest dose groups.
Based on testicular histomorphology indicative of abnormal spermatogenesis in P and Fx
males, the parental and reproductive/developmental toxicity LOAEL and NOAEL are 250 and 50
ppm, respectively. Assuming mean daily doses can be estimated by the mean consumed dose
from weaning to termination of the study, the LOAEL and NOAEL for the F, generation are
22.0 and 4.5 mg/kg/day, respectively; similar doses can be estimated for the parental generation.
Cummings and Hedge (1998) studied the effects of DBA exposure during early pregnancy
in rats. Female Holtzman rats (8/dose group) were administered gavage doses of 0, 62.5, 125, or
250 mg/kg/day on days 1 through 8 of pregnancy. Administration of 500 mg/kg/day induced
moribund behavior and lethality; therefore, dosing was discontinued and these animals were not
further evaluated for reproductive endpoints. Treated animals from the other dose groups were
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sacrificed on day 9 of pregnancy, and the following endpoints were scored: body and reproductive
organ weights; serum levels of progesterone, 17P-estradiol and luteinizing hormone; number of
implantation sites; number of resorptions; corpora lutea and pre-implantation loss. The only
response that was affected was a 170 % increase in serum 17P-estradiol at 250 mg/kg/day. A
second set of females were dosed similarly to the first set of females, sacrificed on gestation day
20, and evaluated for body weight, number of pups, number of resorptions, pup weights, weight
of the placentae and pre-implantation losses. No differences in any of these measures were
observed between treated animals and controls. The authors concluded that DBA had little effect
on female reproduction for the parameters included in the study. They noted that effects on
ovarian function and future fertility were not tested and such tests would be warranted by the
observed increase in serum 17P-estradiol. Based on this study, the NOAEL was 125 mg/kg/day,
based on increase in serum 17P-estradiol; the FEL for acute systemic toxicity is 500 mg/kg/day.
This study is, however, limited by the small size of the dose groups.
There has been considerable interest in the male reproductive tract toxicity of DBA, in
part because DCA is known to be a male reproductive toxicant. Linder and colleagues used a
number of different experimental protocols to study the effects of DBA on spermatogenesis and
the resulting consequences for male fertility, (Linder et al., 1994a; Linder et al., 1994b; Linder et
al., 1995; Linder et al., 1997b). Their studies show that DBA is clearly spermatotoxic following
high-dose single gavage exposures or repeated gavage exposures for longer periods of time (up to
79 days). Effects on spermatogenesis appear to be a sensitive endpoint, since effects are observed
in the absence of other toxicity. In general, histopathological evidence for changes in
seminiferous tubule staging was observed at the lowest doses. Changes in retention of Step 19
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spermatids were generally noted as the earliest effect and occurred following repeated dosing with
10 mg/kg/day, but not 2 mg/kg/day following 31 or 79 daily doses. Significant changes in sperm
count, morphology and motility are generally observed at higher doses than those associated with
early histopathological changes. For example, sperm quality measures (morphology and motility)
were not significantly affected until administration of 50 mg/kg/day for 31 or 79 days (Linder et
al., 1995). Evidence of spermatotoxicity was not necessarily paralleled by changes in male
fertility. For example, even at the high dose of 50 mg/kg/day, which was clearly spermatotoxic,
only marginal (if any) effect on male fertility was observed (Linder et al., 1995). The results of
the studies by Linder and colleagues are supported by those in the two-generation reproductive
toxicity study (CCC, 2001), in which altered spermiation, but no change in male fertility
parameters, was observed. The most sensitive effect of long-term exposure to DBA was the
histopathological effects on the male reproductive tract, with a NOAEL of 2 mg/kg/day (Linder et
al., 1997b).
In contrast to the results by Linder et al. (1994a, 1994b), Vetter et al. (1998) did not find
evidence of spermatotoxicity in another single-dose spermatotoxicity study with sexually-mature
male Crl:CD(SD)BR rats (4-5/dose group) dosed with 0, 600, or 1200 mg/kg DBA in 10 mL/kg
deionized water. No changes in measured sperm parameters (motility, morphology, and cell-
membrane permeability) were reported at either dose; however, mild testes histopathology (the
presence of basophilic bodies) was observed in both dose groups. In the high-dose group, overt
toxicity was observed and included lethargy, irregular gait, decreased feces, ocular discharge,
dyspnea, and abnormal respiratory sounds. No overt toxicity was observed in the low-dose
group. Based on the clinical findings, 1200 mg/kg was considered to be an acute frank effects
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level (FEL). The LOAEL was 600 mg/kg for testes histopathology, and a NOAEL could not be
determined. The reasons for the differences between the findings in this study and those of Linder
et al. (1994) are not known.
Klinefelter et al. (2000), in an abstract, reported the effects of DBA on pubertal
development and adult reproductive function in Sprague-Dawley rats (3 litters/dose) given
drinking water containing 0, 400, 600, or 800 mg/L DBA from gestation day 15 through postnatal
day 98. These drinking water concentrations were chosen to result in doses of 0, 50, 75, and 100
mg/kg/day (personal communication). After weaning, male offspring were exposed to the same
concentrations of DBA in drinking water. Decreased body weight in offspring compared to
controls throughout reproductive development was observed in the high-dose animals.
Histopathology examination revealed seminiferous tubules containing only Sertoli cells in each
dose group, and decreased epididymis weight (the percent decrease was not specified) in the 600
and 800 mg/L dose groups. Fertility of sperm from treated males was decreased by DBA
treatment. The number of implants per corpora lutea in females artificially inseminated with
sperm from treated males decreased from 70% for controls to 49%, 15%, and 15% for the 400,
600, and 800 ppm treatment groups, respectively. Levels of the sperm protein SP22, found to be
highly correlated with fertility, were significantly decreased in all the treatment groups. No
NOAEL for the fertility of sperm of treated males was identified.
In a second recent abstract, Veeramachaneni et al. (2000) exposed male Dutch-belted
rabbits (10/group) to DBA-treated drinking water from gestation day 15 throughout life. The
average daily doses were reported as 0, 0.97, 5.05, and 54.2 mg/kg/day. The fertility of sperm
from 24-week-old males was assessed by artificial insemination of two 6-month-old rabbit does
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per sample of sperm from each male. Conception rates were significantly decreased (p<0.01) in
does inseminated with sperm from males at every dose group. Of the 53 pups born to does
inseminated with sperm from the high-dose males, one pup had both cleft palate and cranioschisis
and two had cranioschisis. At 25 weeks, the offspring were necropsied and no difference from
controls in body weight, anogenital distance or sex organ weights were observed. These data
suggest that the lowest dose tested, 0.97 mg/kg/day, was the LOAEL for decreased male fertility
in rabbits.
In a recent reproductive toxicity study, Klinefelter et al. (2001) confirmed a DBA-related
delay in sexual maturity and effects on sperm quality in Sprague-Dawley rats exposed to DBA. In
this study, the rats were exposed at concentrations of 0, 4, 40, or 400 ppm via drinking water.
Statistically significant, body weight-independent delays in both vaginal opening and preputial
separation were observed in rats exposed to 400 ppm DBA in drinking water (approximately 40
mg/kg/day). The authors stated that pubertal delays were observed only when exposure was
continuous from gestation through adulthood, and not when exposure only occurred from
weaning through adulthood. A related study by the same authors (manuscript in preparation)
found that levels of specific proteins in the sperm membrane (particularly SP 22) were
significantly decreased at 4 and 40 ppm (corresponding to 0.4 and 4 mg/kg/day). As for the
pubertal delays, these changes were only seen in animals that were exposed to DBA via drinking
water from gestation through adulthood, not in animals exposed from weaning through
adulthood. The fertility via artificial (i.e. in utero) insemination was also significantly reduced in
animals exposed to the 400 ppm (40 mg/kg). This study indicates that exposure to DBA in the
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drinking water alters sexual maturity, and sperm quality and fertility, and indicates that effects
may occur at doses lower than previously thought.
No peer-reviewed developmental toxicity studies of DBA are available, but the
developmental toxicity of DBA has been reported in two related abstracts. Narotsky et al. (1996)
studied developmental effects of DBA in CD-I mice dosed by gavage with 0, 0.11, 0.23, 0.46,
0.92, 1.8, 2.8, or 3.7 mmol/kg/day (0, 24, 50, 100, 200, 392, 610, or 806 mg/kg/day) on gestation
days 6-15. Mice were allowed to deliver and their litters were examined on postnatal days 1 and
6. Maternal effects were limited to piloerection and motor depression at the highest dose. There
was delayed parturition at all dose levels, but it is not clear if this effect is adverse. At the highest
dose there was increased prenatal mortality, with only three of nine litters viable at birth.
Increased postnatal mortality was also seen at 2.8 mmol/kg/day (610 mg/kg/day), and 3.7
mmol/kg/day (806 mg/kg/day). Decreased pup weight was observed at 3.7 mmol/kg/day on
postnatal day 1. Pup weight was also decreased on postnatal day 6 at 2.8 mmol/kg/day. Short,
kinked or absent tails were observed at the two highest doses. Based on these results, the authors
concluded that DBA was a developmental toxicant.
In a second published abstract, DBA was administered to CD-I mice by gavage in distilled
water on gestation days 6-15, at doses of 0, 50, 100, or 400 mg/kg/day (Narotsky et al., 1997b).
Maternal toxicity was not observed. Litters were removed by cesarean section on gestation day
17, and half of the fetuses in each litter were examined for skeletal defects and the other half for
visceral effects. There was no effect on prenatal survival, fetal weight or skeletal development.
Hydronephrosis was noted at 100 and 400 mg/kg/day, as well as renal agenesis (small kidneys) at
400 mg/kg/day. Based on the summary data provided in these abstracts, the NOAEL would be
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50 mg/kg/day and the LOAEL for fetal kidney malformations would be 100 mg/kg/day.
In addition to the developmental studies in animals, the developmental toxicity of DBA
was evaluated in whole embryo culture studies. Hunter et al. (1996) reported on the
developmental effects of DBA in whole embryo culture. DBA induced malformations in the
embryos at sub-lethal doses. Andrews et al. (1999) evaluated the developmental effects of DBA
in rat embryos. Gestational day 9 embryos were exposed to 10 to 400 |iM DBA for 48 hours.
Concentrations of 200 |iM or greater DBA resulted in developmental effects.
In another whole-embryo culture study designed to test mechanisms of haloacetic acid-
induced dysmorphogenesis, Ward et al. (2000) reported that DBA caused only a limited induction
of sub-Gl cell-cycle events (an increase in hypodiploid cells or cell debris resulting from DNA
breakage during apoptosis). However, based on the more significant effects of BCA and DCA
treatment on the induction of apoptosis, the authors suggested the haloacetic acids might induce
embryo dysmorphogenesis through their ability to increase apoptosis.
Negative results were obtained in a non-mammalian screening assay regarding the
developmental toxicity of DBA. Gardner and Toussant (1999) evaluated developmental toxicity
of DBA in the frog embryo teratogenesis assay - Xenopus (FETAX) (a 96-hour toxicity test),
with and without metabolic activation. Endpoints evaluated were embryolethality (LC50),
embryonic malformations (ECS0), minimum concentration to inhibit growth (MCIG), and a
teratogenicity index (TI - the ratio of the LCS0 to the EC50). The FETAX assay is considered to
be a reliable developmental toxicity screening assay; Dawson and Bantle (1987) have estimated
that its predictive accuracy for identifying known mammalian or human developmental toxicants
approaches or exceeds 85%. Under the conditions of this study, DBA did not exhibit teratogenic
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potential. Further, malformations did not appear to increase in severity or prevalence with
increasing DBA concentrations, with or without metabolic activation. Although negative result
was obtained in one study, the overall whole-embryo testing data provide support for the
developmental toxicity of the brominated acetic acids observed in vivo.
Systemic Effects
The liver toxicity, immunotoxicity, and neurotoxicity of DBA have been evaluated.
Parrish et al. (1996) exposed male mice to DBA-treated drinking water for 21 days. Estimated
doses based on default water-intake values for mice were 0, 25, 125, or 500 mg/kg/day.
Increased liver weight was observed beginning at 125 mg/kg/day and supported by evidence of
oxidative stress at the same dose; the NOAEL was 25 mg/kg/day. In an immunotoxicity study
(NTP, 1999), female mice were given DBA in their drinking water for 28 days. The resulting
DBA doses were not reported by the study authors, but based on body weight and water-
consumption data, doses were estimated for each of four sub-studies conducted. Absolute and
relative liver weights were increased beginning at 14 mg/kg/day and increased in a dose-
dependent fashion. Liver weight changes were not chosen as the critical effect for this study due
to the absence of histopathology or clinical chemistry confirming that the observed liver-weight
increases were adverse. Analogy to other haloacetic acids suggests that DBA is likely to induce
adverse liver effects (histopathology or clinical chemistry changes) at sufficiently high doses, but
none of these effects were observed in this study.
The immunotoxicity of DBA administered in drinking water has been evaluated in four
studies in mice exposed to drinking water containing 0, 125, 250, 500, 1000, or 2000 mg/L DBA
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for 28 days (NTP, 1999). A number of different end points were assessed, including thymus and
spleen weights, number and type of spleen cells, macrophage activation, natural killer (NK) cell
activity, and specific and general IgM antibody-forming responses. The most sensitive and
reliable measure was a decrease in spleen IgM antibody-forming cell responses, representing a
clear decrease in immune system function, accompanied by an increase in the number of spleen
macrophages. The LOAEL and NOAEL for these endpoints were approximately 70 and 38
mg/kg/day, respectively.
Phillips et al. (2002, published abstract) examined the neurobehavioral toxicity of DBA in
adolescent (28-day-old) male and female F344 rats (12/sex/dose) given DBA in drinking water at
concentrations of 0, 200, 600, or 1500 mg/L (mean doses calculated by the authors as 0, 20, 72,
and 161 mg/kg/day) for 6 months. In both sexes, body weight was significantly depressed in the
highest dose group but overall health status was unaltered. A neurobehavioral test battery was
administered to all animals at 1, 2, 4, and 6 months. Dose-dependent neuromuscular toxicity,
characterized by mild gait abnormalities, hypotonia, and decreased forelimb and hindlimb grip
strength, was observed in both sexes. Sensorimotor responsiveness, as measured by responses to
a tail pinch and click, was reduced at all doses but did not progress with continued exposure to
DBA. Decreased motor activity was noted in both sexes in the high-dose group, whereas a chest
clasping response was only observed in high-dose females. Neuropathologic examination
revealed significant myelin fragmentation, axonal swelling, and axonal degeneration in the lateral
and ventral areas of the spinal cord white matter in the high-dose group. In the mid- and high-
dose groups, small numbers of swollen, eosinophilic or faintly basophilic, and occasionally
vacuolated neurites were observed in the spinal cord gray matter, and appeared to represent
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axonal degeneration. Neuropathologic examination has not yet been conducted in the low-dose
group. No treatment-related neuropathology was noted in the eyes, peripheral nerves, peripheral
ganglia, or brain. Based on neurobehavioral abnormalities, the LOAEL was 20 mg/kg/day, the
lowest dose tested, and a NOAEL could not be determined
Carcinogenicity
No complete reports of bioassays with DBA have been published. In published abstracts,
So and Bull (1995) reported that DBA induces aberrant crypt foci in the colon of rats, and
Stauber et al. (1995) reported that DBA induces liver tumors in mice. DBA has also provided
nearly uniformly positive results in the genotoxicity assays conducted. Positive results have been
reported in S. typhimurium assays (Giller et al., 1997; Kohan et al., 1998; NTP 2000c) and assays
for DNA damage repair (Giller et al, 1997; Mayer et al., 1996). DBA has been shown to induce
oxidative DNA damage (Austin et al., 1996; Parrish et al., 1996). On the other hand, no
induction of micronuclei was reported in a newt larvae system (Giller et al., 1997). The
clastogenicity of DBA has not been reported in other assays using a standard protocol, but DBA
has been reported to be co-clastogenic (Sasaki and Kinae, 1995). As a whole, these data support
the conclusion that DBA is genotoxic.
In the absence of full cancer bioassay data, DBA is classified as Group D: Not Classifiable
under the 1986 Cancer Risk Assessment Guidelines. Under EPA's Draft 1999 Guidelines for
Carcinogen Risk Assessment, there is suggestive evidence of carcinogenicity of DBA, but the
evidence is not sufficient to assess human carcinogenic potential. The existing weight of the
evidence in support of the carcinogenicity of DBA includes its structural analogy to the
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demonstrated rodent tumorigen DCA, strong evidence for genotoxicity, and positive preliminary
results in cancer studies reported in published abstracts (So and Bull, 1995; Stauber et al., 1995).
Thus, the animal data raise a concern for potential carcinogenic effects, but are not sufficient for
reaching a conclusion with regard to human carcinogenicity. A 2-year NTP toxicity and
carcinogenicity study with DBA is ongoing (NTP, 2000a).
Basis for RfD and MCLG
The data on DBA are insufficient for the derivation of an RfD, because there are no
systemic toxicity studies of sufficient duration. Similarly, there are insufficient data to assess the
carcinogenic potential of DBA. In the absence of adequate data on the noncancer or cancer
effects of DBA, no MCLG is proposed.
Children's Risk in Relation to the MCLG
The data are mixed as to whether children may be more sensitive than adults to the effects
of DBA. Christian et al. (2001b) found no effects in the pups of rats continuously administered
DBA in drinking water during premating and mating (adult males and females) and gestation and
lactation (females). Similarly, pups given DBA in drinking water for one week following weaning
did not exhibit any treatment-related effects. Christian et al. (2001b) also analyzed the
distribution of DBA in various tissues. At high drinking-water concentrations, DBA was detected
in the placenta, amniotic fluid, and fetal plasma, indicating that it readily crossed the placenta and
distributed to the fetus, although it did not appear to bioaccumulate. In the two-generation
reproductive toxicity study (CCC, 2001), a dose-related increase in altered sperm
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histomorphology and histopathology of the male reproductive tract was observed in both the
parental and Fx generations, indicating that the male reproductive tract is a target organ in both
adult and developing rats. However, the incidence of affected animals was higher in the parental
than the Fx generation, and thus developing males did not appear to be more sensitive than adults.
In contrast, in two published abstracts, (Narotsky et al., 1996; 1997b) DBA was reported
to induce developmental toxicity when administered by gavage to CD-I mice. Differences in the
results between these studies and those of Christian et al. (2001b) and the Chlorine Chemistry
Council (2001) may have been due to differences in the route of dosing (gavage versus drinking
water), species sensitivity (mouse versus rat), dosing regime, or other study details not reported in
the published abstracts. These data are supported by recent reports by Klinefelter et al. (2001;
manuscript in preparation), that reported a DBA-related delay in sexual maturity and effects on
sperm quality in Sprague-Dawley rats exposed to DBA in drinking water. These effects were
independent of changes in body weight, and were seen when exposure was from gestation
through puberty, but not when exposure only occurred from weaning through puberty. Pubertal
delays occurred at doses as low as 400 ppm DBA in drinking water (approximately 40
mg/kg/day), while changes in sperm membrane proteins were observed at doses as low as 4 ppm
(0.4 mg/kg/day). The risk assessment implications of this study are currently being evaluated.
Veeramachaneni et al. (2000) reported in an abstract that exposure of rabbits in utero
from gestation day 15 to 24 weeks of age reduced the fertility of sperm from treated males. The
lowest dose tested, 0.97 mg/kg/day was the LOAEL. This LOAEL for fertility changes was 10-
fold lower than the LOAEL of 10 mg/kg/day reported in Linder et al. (1997b) for altered
histopathology.
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2.4. BROMATE
Bromate (Br03") is formed in water following disinfection through ozonation of water
containing bromide ion. In laboratory studies, the rate and extent of bromate formation depends
on the ozone concentration used in disinfection, pH and contact time.
2.4.1. Developmental/Reproductive Effects
Limited data are available on the reproductive or developmental effects of bromate. No
standard reproductive, developmental, or multigeneration studies are available for bromate.
Kurokawa et al. (1990) reported the results of a 5-generation study in mice and an 8-generation
study in mice, in which the animals were fed bread made from flour treated with 14-100 ppm
potassium bromate; information on the study designs was not presented. No effects on behavior,
weight gain, reproductive performance or histological abnormalities were reported (Kurokawa et
al., 1990). However, since most potassium bromate added to flour is converted to bromide
during the bread-baking process (Kurokawa et al., 1986b), it is unlikely that the animals in these
multigeneration studies were actually exposed to bromate.
In a screening study conducted for NTP, Wolf and Kaiser (1996) evaluated the potential
reproductive and developmental toxicity of sodium bromate in Sprague-Dawley rats following
oral administration in the drinking water at concentrations of 0.25 ppm (2.6 mg/kg/day), 80 ppm
(9.0 mg/kg/day), or 250 ppm (25.6 mg/kg/day) over a 35-day period. (Equivalent bromate ion
doses are 2.2, 7.7, and 22 mg Br03-/kg/day.) Two groups of female rats were treated. Group 1
females (10/group) were dosed from study day 1 to 34 to test for effects during conception and
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early gestation. Group 2 females (13/group) were dosed from gestation day 6 to postnatal day 1
to test for effects during late gestation and birth. Male rats (10/group) were cohabited with
Group 2 females for 5 days prior to dosing (study days 1-5) and were then dosed from study day
6 to day 34/35. Females in Group 2 were allowed to litter and the pups were observed through
postnatal day 5. However, there is no indication of the developmental endpoints that were
evaluated in these pups or if any effects were observed. Treated males in the 250 ppm dose group
demonstrated a statistically significant decrease (18%) in epididymal sperm density. All other
endpoints evaluated were comparable between controls and treated groups. Female reproductive
function was not adversely affected. There were no treatment-related gross or microscopic
changes in the kidney, liver, spleen, testis or epididymis. These results indicated that sodium
bromate treatment did not produce any adverse signs of general toxicity in any of the dose levels
tested; a maximum tolerated dose (MTD) was not reached. Based on changes in sperm density,
this study identified a NOAEL of 80 ppm (7.7 mg Br037kg/day) and a LOAEL of 250 ppm (22
mg/Br03Vkg/day).
2.4.2. Systemic Toxicity
A number of cases of acute bromate intoxication have been reported in humans (both
children and adults) following accidental or suicidal ingestion of permanent hair-wave neutralizing
solutions, which usually contain either 2% potassium bromate or 10% sodium bromate. The case
studies suggest that there are no differences between children and adults in either the target organ
or the effective dose following acute oral exposure to bromate. No epidemiological studies were
located on noncarcinogenic or carcinogenic effects of bromate exposure in humans.
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Several authors report the effects of acute oral exposure in children to potassium bromate
following accidental ingestion of hair home permanent neutralizing solution (Benson, 1951;
Parker and Barr, 1951; Quick et al., 1975; Gradus et al., 1984; Warshaw et al., 1985; Lue et al.,
1988; Mack, 1988; Lichtenberg et al., 1989; Watanabe et al., 1992). The age of the children,
when reported, ranged from 17 months (Gradus et al., 1984) to 6 years (Quick et al., 1975).
When estimated, doses ranged from 20 mg Br03-/kg (Watanabe et al., 1992) to 1,000 mg Br03-
/kg (Lue et al., 1988). In all cases, the initial symptoms appeared to include abdominal pain,
vomiting or other gastrointestinal effects. Central nervous system (CNS) effects such as sedation,
lethargy, and CNS depression appeared to be early symptoms of bromate poisoning after doses of
about 70 mg/kg or higher (Parker and Barr, 1951; Warshaw et al., 1985; Lue et al., 1988;
Lichtenberg et al., 1989). Irreversible deafness is also an effect of bromate exposure (Quick et al.,
1975; Gradus et al., 1984); one review of bromate ototoxicity found that deafness occurred in 18
of 31 cases, usually within 4-16 hours of exposure (Matsumoto et al., 1980).
Kidney effects were frequently observed in children following acute exposure; although
there is not a clear relationship between dose and the development of renal effects. One review
of bromate kidney toxicity found that renal failure occurred in 26 of 31 reported cases
(Matsumoto et al., 1980). Anuria persisting for several days or longer was observed following
exposure to 20 mg/kg potassium bromate (Quick et al., 1975) up to doses of 1,000 mg Br03-/kg
(Lue et al., 1988). In contrast, children ingesting 20 mg Br03-/kg (Watanabe et al., 1992) and
children ingesting 230-460 mg Br03-/kg (Lichtenberg et al., 1989) did not demonstrate any renal
effects. Histological examination of renal biopsies from children with renal effects indicated
interstitial edema, interstitial fibrosis, tubular atrophy (Quick et al., 1975), and epithelial
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separation of the proximal tubules (Watanabe et al., 1992). Glomeruli were not affected.
Although there are fewer reports of acute oral exposure to bromate in adults (Matsumoto
et al., 1980; Kuwahara et al., 1984; Kutom et al., 1990; Hamada et al., 1990), the symptoms of
toxicity appear to be similar to those observed in children. When reported, the doses ingested
ranged from 100-150 mg Br03-/kg (Matsumoto et al., 1980) to 500 mg K Br03/ kg (Kuwahara et
al., 1984). In all cases, the first symptoms to appear were gastrointestinal, including nausea,
vomiting, diarrhea and abdominal pain. Hearing loss was reported by three authors (Matsumoto
et al., 1980; Kuwahara et al., 1984; Hamada et al., 1990). Anuria and renal failure were also
reported (Kuwahara et al., 1984; Kutom et al., 1990; Hamada et al., 1990). The amount of time
required to recover renal function varied from 7 days (Kutom et al., 1990) to 5 weeks (Hamada et
al., 1990), and in two cases, renal function was never restored (Kuwahara et al., 1984).
Several subchronic or chronic studies in animals indicate that the kidney is the primary
target organ following long-term oral exposure to bromate. Following a 13-week exposure of
rats to a dose of 63 mg Br03-/kg/day (Kurokawa et al., 1990), the following non-neoplastic
effects were observed: inhibition of body-weight gain; significant increases in several serum
parameters, including blood urea nitrogen (BUN); and droplets of various sizes and regenerative
changes in the renal tubules. Similar effects were observed in chronic studies of oral bromate
exposure (Nakano et al., 1989; Kurokawa et al., 1986b; DeAngelo et al., 1998). The following
non-neoplastic effects have been reported following long-term exposure: increased BUN;
increased severity of nephropathic changes; degenerative and necrotic kidney lesions including
hyaline casts in the tubular lumen, hyaline droplets, eosinophilic bodies, and brown pigments in
the tubular epithelium; and urothelial hyperplasia of the transitional epithelium of the renal pelvis.
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No non-neoplastic effects have been reported in tissues other than the kidney.
DeAngelo et al. (1998) observed statistically significant increases in relative liver weight,
absolute and relative kidney weight, absolute and relative thyroid weight, and relative spleen
weight in rats treated with 28.7 mg/kg/day in drinking water. Non-neoplastic kidney lesions
observed in rats included a significant dose-dependent increase in the incidence of urothelial
hyperplasia at doses of 6.1 mg/kg/day and higher, foci of mineralization of the renal papilla and
eosinophilic droplets in the proximal tubule epithelium. There were no other treatment-related
non-neoplastic effects observed in any other tissue examined. Based on urothelial hyperplasia in
male rats, this study identifies a NOAEL of 1.1 mg Br03-/kg/day and a LOAEL of 6.1 mg
Br03-/kg/day.
2.4.3. Carcinogenicity
There are no epidemiology data regarding the carcinogenic potential of bromate.
DeAngelo et al. (1998) administered potassium bromate to male rats and male mice in drinking
water for 100 weeks. Statistically significant, dose-dependent increases in tumors of the kidney,
thyroid, and tunica vaginalis testis (mesotheliomas) were observed. This study contributes to the
weight of evidence for the potential human carcinogenicity of bromate and confirms the study by
Kurokawa et al. (1986a,b), which reported renal tumors in male rats.
The evidence is too limited to reach a conclusion about any mode of action (EPA, 2001b).
The genotoxicity of bromate has been evaluated in a variety of in vitro and in vivo systems, with
consistently positive results. Oxidative stress may play a role in the formation of kidney tumors
induced by bromate. Two mechanisms of bromate-mediated DNA damage have been proposed:
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direct interaction with DNA following GSH activation, and indirect damage via lipid peroxides.
The data suggest that, in the intact kidney, bromate induces DNA damage through lipid
peroxidation at toxic doses, rather than via a direct mechanism (Chipman et al., 1998). However,
the overall evidence is insufficient to establish lipid peroxidation and free radical production as the
key events responsible for the induction of kidney tumors, and the data are insufficient to
implicate any single mechanism for the production of thyroid and testicular tumors (EPA, 2001b).
Some evidence suggests that cell proliferation related to a2u-globulin plays a role in enhancing
renal carcinogenesis by bromate (Umemura et al., 1993), but the observation of kidney tumors in
female rats, and the observation of testicular and thyroid tumors in male rats suggests that a2u-
globulin is not the primary mechanism of bromate carcinogenicity (EPA, 2001b).
Under EPA's Guidelines for Carcinogen Risk Assessment (EPA, 1986), bromate would be
classified as Group B2: Probable Human Carcinogen based on no evidence in humans and
adequate evidence of carcinogenicity in male and female rats (EPA, 2001b). Under the Draft
Guidelines for Carcinogen Risk Assessment (EPA, 1999), bromate is likely to be carcinogenic to
humans by the oral route of exposure (EPA, 2001b). Insufficient data are available to evaluate
the human carcinogenic potential of bromate by the inhalation route. Although no
epidemiological studies or studies of long-term human exposure to bromate are available, bromate
is carcinogenic to male and female rats following exposure in drinking water. This weight-of-
evidence conclusion is based on sufficient experimental findings that include the following: tumors
at multiple sites in rats, tumor responses in both sexes, and evidence of mutagenicity including
point mutations and chromosomal aberrations in vitro.
Cancer risk estimates were derived from the DeAngelo et al. (1998) study by applying the
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one-stage Weibull model for the low-dose linear extrapolation. Bromate was administered to
male F344 rats or B6C3F1 mice in drinking water at concentrations of 0, 0.02, 0.1, 0.2, and 0.4
g/L or 0, 0.08, 0.4, and 0.8 g/L, respectively, for 108 weeks. Using a time-to-tumor model and a
Monte Carlo analysis to sum across the three tumor sites, the upper-bound cancer potency for
bromate ion is estimated to be 0.7 per mg/kg/day. Assuming a daily water consumption of 2 L for
a 70 kg adult, lifetime risks of 10"4, 10"5, and 10"6 are associated with bromate concentrations in
water of 5, 0.5, and 0.05 |ig/L, respectively. This estimate of cancer risk from the DeAngelo et
al. study is similar to the risk estimate derived from the Kurokawa et al. (1986a) study in the 1994
proposed rule (EPA, 1994a).
2.4.4. Basis for the RfD and MCLG
The Agency based the bromate RfD on urothelia hyperplasia in male rats in the DeAngelo
et al. (1998) study, with a NOAEL of 1.1 mg Br03-/kg/day and a LOAEL of 6.1 mg
Br03-/kg/day (EPA, 2001b). Kidney effects have also been seen in children and adults following
accidental high-dose exposure, indicating animal to human concordance in target. An UF of 10
was applied for extrapolating from animals to humans, and another factor of 10 was used for
variability in sensitivity among members of the human population. A factor of 3 was used to
account for some deficiencies in the database. The bromate database consists of chronic and
subchronic studies in rats and mice and a screening level reproductive/developmental study in rats.
The database is missing developmental toxicity studies in two species and a standard
multigenerational study. This results in a total UF of 300. The resulting RfD for bromate is 0.004
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mg/kg/day. This RfD corresponds to a DWEL of 0.14 mg/L, assuming an adult tap water
consumption of 2 L/day for a 70 kg adult.
The MCLG for bromoform is zero, based on its likely human carcinogenicity.
2.4.5. Children's Risk in Relation to the MCLG
The developmental toxicity of bromate has not been adequately evaluated, so no
conclusions regarding the sensitivity of the fetus can be drawn. Limited evidence suggests that
bromate may be a male reproductive toxicant, although at a higher dose than that which results in
kidney toxicity. No data were identified that describe the effects of in utero or neonatal exposure
to bromate. Case reports on the effects of acute oral exposure to bromate suggest that systemic
toxicity is similar in children and adults. No data were identified regarding age-related differences
in absorption, distribution, metabolism or excretion of bromate. The bromate MCLG is zero,
based on its likely human carcinogenicity. The mode of action for bromate is unclear; however,
evidence of genotoxicity led to use of a linear low-dose extrapolation. The Agency believes that,
since there are no data suggesting that children are more sensitive to bromate, the MCLG of zero
is protective of both children and adults.
2.5. CHLORITE/CHLORINE DIOXIDE
Chlorite and chlorine dioxide are evaluated together in this assessment for children's risk.
It is likely that the studies conducted with chlorite, the predominant degradation product of
chlorine dioxide, are relevant to characterizing the toxicity of chlorine dioxide. In addition,
studies conducted with chlorine dioxide may be relevant to characterizing the toxicity of chlorite.
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Chlorine dioxide is fairly unstable and rapidly dissociates to form predominantly chlorite
and chloride, and to a lesser extent, chlorate. There is a ready interconversion among these
species in water (before administration to animals) and in the gut (after ingestion) (EPA, 1994c).
Therefore, what exists in water or stomach is a mixture of these chemical species (i.e., chlorine
dioxide, chlorite and chlorate), and possibly the products of their reactions with the
gastrointestinal contents. As a result, the toxicity data for chlorine dioxide or chlorite are
considered applicable for assessing the toxicity of the other. In the study descriptions below,
concentrations in drinking water are provided as reported by the study authors, but all doses are
converted to concentrations of chlorite or chlorine dioxide.
2.5.1. Developmental/Reproductive Effects
There is a large database on the developmental and reproductive toxicity of chlorite and
chlorine dioxide. Two epidemiologic studies examined the relationship between water disinfected
with chlorine dioxide and developmental toxicity (Kanitz et al., 1996; Tuthill et al., 1982). Tuthill
et al. (1982) retrospectively compared infant morbidity and mortality data for a community using
chlorine dioxide as a drinking-water disinfectant in the 1940s with a community using
conventional drinking-water chlorination treatment. No effect was observed on
fetal, neonatal, postneonatal or infant mortality; however, the incidence of newborns judged
premature by physician assessment was significantly higher in the chlorine dioxide-treated
community. EPA (1994d) concluded that there was no increase in the proportion of premature
infants when the age of the mother was controlled and that there was a greater postnatal weight
loss in infants from the exposed community.
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As summarized in Section 2.1 on Chlorinated Drinking Water, Kanitz et al. (1996)
conducted an analysis of 676 births from hospital records of two Italian cities that used
chlorinated surface water, surface water treated with chlorine dioxide, or well water that was not
treated. The authors found that the frequency of newborns with short body length, small cranial
circumference, and low birth weight was increased for mothers over 30 years of age who were
served by chlorinated water supplies. The frequency of newborns with neonatal jaundice was
increased for births to mothers exposed to water treated with chlorine dioxide.
Six oral studies have examined the developmental toxicity of chlorite in animals, including
four studies in rats (CMA, 1996; Couri et al., 1982; Mobley et al., 1990; Suh et al., 1983) and one
study each in mice (Moore et al., 1980) and rabbits (Harrington et al., 1995b). The Chemical
Manufacturers Association (CMA) conducted a two-generation study examining
the developmental neurotoxicity, reproductive, and hematologic effects of rats exposed to sodium
chlorite (CMA, 1996). Males were exposed throughout mating, and the females were exposed
during mating, pregnancy, and lactation. Thirty male and 30 female Sprague-Dawley rats
received drinking water containing 0, 35, 70, or 300 mg/L sodium chlorite for 10 weeks and were
then mated. Based on intakes calculated by the study authors, doses for the P animals were 0, 3,
5.6, or 20, and 0, 3.8, 7.5, or 28.6 mg/kg/day chlorite for males and females, respectively. For the
Fx animals, doses were 0, 2.9, 5.9, or 22.7 mg/kg/day chlorite for the males and 0, 3.8, 7.9, or
28.6 mg/kg/day chlorite for the females. At 300 mg/L, effects observed in at least one generation
and sex included decreased pup survival and pup body weight, decreased absolute and relative
liver weight, delayed sexual development, decreased red blood cell parameters, and several
neurotoxic effects (e.g., decreased brain weight and increased incidence of abnormal righting
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reflex). Effects at 70 mg/L included decreased absolute and relative liver weights. A significant
decrease in maximum response to an auditory startle stimulus was noted in the 70 and 300 mg/L
groups on postnatal day 24, but not on postnatal day 60. The NOAEL in this study was 35 mg/L
(2.9 mg/kg/day chlorite) and the LOAEL was 70 mg/L (5.9 mg/kg/day chlorite) based on lowered
auditory startle amplitude and altered liver weights in two generations of rats.
Mobley et al. (1990) exposed groups of 12 female Sprague-Dawley rats to 0, 20 or 40
mg/L chlorite in drinking water (estimated as 0, 3, or 6 mg/kg/day) for 10 days prior to mating
with unexposed males, and during gestation and lactation until postnatal days 42-53. Significant
decreases in exploratory activity were observed in the 6 mg/kg/day group on postnatal days 36-39
but not on days 40-41. In the 3 mg/kg/day group, significant decreases in exploratory activity
were observed on days 36 and 37 but not on days 38-41. Free T4 was significantly increased in
pups at 6 mg/kg/day. Based on a review of the results of this study relative to the findings of the
newer developmental studies in the chlorite database, EPA (2000c) concluded that the NOAEL
for neurobehavioral effects in this study was 3 mg/kg/day and the LOAEL was 6 mg/kg/day.
Suh et al. (1983) exposed groups of six to nine female Sprague-Dawley rats to 0, 1, or 10
mg/L chlorite in drinking water (estimated as 0, 0.1, or 1 mg/kg/day) for 2.5 months prior to
mating with unexposed males and during gestational days 0-20. No significant alterations in
maternal body weight gain, number of implants, resorptions, dead fetuses, fetal body weights, or
incidence of skeletal abnormalities were observed at any dose. Crown-rump length was
significantly higher in the 1 mg/kg/day group compared with controls, but the difference was small
and probably not biologically significant. A NOAEL of 1 mg/kg/day for developmental toxicity
was identified in this study.
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In another developmental study, Couri et al. (1982) exposed groups of 7-13 pregnant
Sprague-Dawley rats to 0%, 0.1%, 0.5%, or 2% sodium chlorite (0%, 0.01%, 0.4%, or 1.5%
chlorite; estimated as 0, 70, 440, or 610 mg/kg/day chlorite) in drinking water during gestational
days 8-15. Another group of four pregnant rats received daily gavage doses of 200 mg/kg/day
sodium chlorite on gestational days 8-15. One hundred percent mortality was observed in the rats
receiving this gavage dose, while no deaths were observed in the rats received sodium chlorite via
drinking water. In the rats receiving sodium chlorite via drinking water, irregular blood cells,
ruptured cells and hemolysis were observed in the 610 mg/kg/day group. Significant decreases in
crown-rump length and an increase in the number of resorbed and dead fetuses in litters delivered
on gestational day 22 were observed in all dose groups. Postnatal growth and the incidences of
soft tissue and skeletal malformations were not adversely affected in this study. This study
identified a frank effect level (FEL) of 70 mg/kg/day for resorbed and dead fetuses and decreases
in crown-rump length.
Moore et al. (1980) treated groups of pregnant A/J mice with 0 or 100 mg/L sodium
chlorite in drinking water (75 mg/L chlorite, corresponding to approximately 22 mg/kg/day
chlorite) throughout gestation and lactation. A decrease in the conception rate was observed in
the chlorite group. Pup growth was adversely affected, as shown by significant decreases in
average pup weaning weight and birth to weaning growth rate. This study identified a LOAEL of
22 mg/kg/day for developmental effects.
Harrington et al. (1995b) exposed groups of 16 New Zealand white rabbits to sodium
chlorite in drinking water at concentrations of 0, 200, 600, or 1,200 mg/L (0, 10, 26 or 40
mg/kg/day chlorite) on gestation days 7-20. No treatment-related effects on pregnancy incidence,
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number of malformations, number of preimplantation losses, fetal sex ratio, number of live fetuses
or fetal visceral or structural abnormalities were observed at any dose. In the 26 and 40
mg/kg/day groups, an increased incidence of minor skeletal abnormalities and skeletal variants
was noted. The NOAEL in this study was 10 mg/kg/day and the LOAEL was 26 mg/kg/day,
based on decreased fetal weight and delayed skeletal ossification, and decreased food and water
consumption and decreased body weight gain in the dams.
Several studies investigated the reproductive toxicity of chlorite in drinking water. Carlton
and Smith (1985) and Carlton et al. (1987) exposed groups of 12 male rats to sodium chlorite in
drinking water for 56 days prior to mating and throughout a 10-day mating period; they also
exposed groups of 24 female rats to the same sodium chlorite concentrations for 14 days prior to
mating and throughout gestation and lactation. No significant alterations in body weight gain,
fertility rates, litter-survival rates, median day of eye opening, or median day of observed vaginal
patency were noted. No alterations were seen in the reproductive tract tissues, hematologic
parameters, testis, epididymis, caudal epididymis weights, or percentage of abnormal sperm. The
only effects noted were significant decreases in T3 and T4 hormone levels. In follow-up studies,
groups of 12 male rats received sodium chlorite in drinking water for 72-76 days and a
statistically significant increase in abnormal sperm at 100 mg/L sodium chlorite (70 mg/L chlorite,
or approximately 7.5 mg/kg/day) was noted; the NOAEL was 0.75 mg/kg/day chlorite.
Reproductive function was not assessed in this study.
Five oral studies investigated the developmental toxicity of chlorine dioxide in rats
(Mobley et al., 1990; Orme et al., 1985; Suh et al., 1983; Taylor and Pfohl, 1985; Toth et al.,
1990). Several of these studies tested only one dose. Orme et al. (1985) exposed groups of
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female Sprague-Dawley rats to 0, 2, 20 or 100 mg/L chlorine dioxide in drinking water (estimated
at 0, 1, 3 or 14 mg/kg/day) for 2 weeks prior to mating and throughout gestation and lactation.
Additionally, groups of 5-day-old Sprague-Dawley pups received gavage doses of 0 or 14
mg/kg/day chlorine dioxide on postnatal days 5-20. In the 14 mg/kg/day gavage group, activity
was significantly decreased on postnatal days 18-19; on days 15-17 and 20, activity levels were
similar to controls. In the 14 mg/kg/day drinking-water group, there was a significant decrease in
T3 and T4 levels; in all groups there was a significant correlation between T4 levels and
locomotor activity. T4 levels were not significantly altered in the chlorine dioxide-exposed dams.
ANOAEL of 3 mg/kg/day and a LOAEL of 14 mg/kg/day for neurobehavioral effects (decreased
T3 and T4 levels and delayed development) were identified in this study.
Suh et al. (1983) exposed groups of six to eight female Sprague-Dawley rats to 0, 1, 10,
or 100 mg/L chlorine dioxide in drinking water (approximately 0, 0.1, 1, or 10 mg/kg/day) for 2.5
months prior to mating with unexposed males and during gestational days 0-20. There was a
statistically significant trend for decreasing number of implants per litter and number of live
fetuses per dam. Total fetal weights and male fetal weights were significantly increased in the 10
mg/kg/day group; crown-rump length was not significantly affected. Incidences of skeletal
abnormalities did not significantly differ between groups. This study identified a NOAEL of 1
mg/kg/day and a LOAEL of 10 mg/kg/day for developmental effects.
Mobley et al. (1990) exposed groups of 12 female Sprague-Dawley rats to 0 or 100 mg/L
chlorine dioxide in drinking water (estimated at 0 or 14 mg/kg/day) for 10 days prior to mating
with unexposed males and during the gestation and lactation periods (through postconception day
41). At birth, the litter weight of the chlorine dioxide-exposed group was significantly lower than
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that of controls. Chlorine dioxide exposure also significantly decreased exploratory activity on
postconception days 36-39, but not on days 40-41. This study identified a LOAEL of 14
mg/kg/day for decreased litter weight and decreased exploratory activity.
In another developmental study, Taylor and Pfohl (1985) exposed groups of 13-16 female
Sprague-Dawley rats to 0 or 100 mg/L chlorine dioxide in drinking water (estimated at 0 or 14
mg/kg/day) for 14 days prior to breeding and throughout gestation and lactation. In another
phase, groups of male pups from unexposed dams were administered 0 or 14 mg/kg/day chlorine
dioxide via gavage from postnatal days 5 to 20. No effects on maternal or pup body weights were
observed in the exposed group, but a significant decrease in whole-brain weight was observed in
the 21-day-old offspring of dams receiving 14 mg/kg/day in drinking water. A nonsignificant
decrease in locomotor activity was observed in the 14 mg/kg/day offspring assessed at 10-20 days
of age, and a significant decrease in exploratory activity was observed at this dose at 60 days of
age. In the gavage-dose group, significant decreases in body weight, whole-brain and forebrain
weights, forebrain DNA content and decreases in home cage and wheel-running activity were
observed in the pups. The LOAEL for neurobehavioral effects, decreased brain weight and cell
number in this study was 14 mg/kg/day, based on both drinking water and gavage dosing.
In a neurodevelopmental toxicity study by Toth et al. (1990), groups of four male and four
female pups per litter received gavage doses of 0 or 14 mg/kg/day chlorine dioxide on postnatal
days 1-20. Forebrain weights were significantly lower in the chlorine dioxide-exposed pups on
postnatal days 21 and 35. No gross lesions, and loss of myelin in the forebrain, cerebellum or
brainstem were observed in the chlorine dioxide-exposed pups. Thus, a LOAEL of 14 mg/kg/day
for altered brain development was identified in this study.
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Reproductive and developmental toxicity of chlorine dioxide were studied by Carlton et al.
(1991). Daily gavage doses of 0, 2.5, 5, or 10 mg/kg/day chlorine dioxide in deoionized water
were administered to groups of 12 male and 24 female Long-Evans rats prior to mating and
throughout the mating period. The females continued to receive the doses throughout gestation
and lactation. No effects were noted on mortality, clinical signs, fertility rates, sperm parameters,
length of gestation, prenatal deaths, mean litter size or mean pup weights. A NOAEL of 10
mg/kg/day for reproductive effects was noted from this study.
The observed neurodevelopmental effects (CMA, 1996; Mobley et al., 1990; Orme et al.,
1985; Taylor and Pfohl, 1985; Toth et al., 1990) may have been related to decreased T3 and T4
levels (Mobley et al., 1990; Orme et al., 1985). Hypothyroidism, defined by increased serum
levels of TSH and decreased levels of T3 and T4, has been associated with neurodevelopmental
delay in both humans and rats. In humans, congenital hypothyroidism, or cretinism, is
characterized by long-term effects on behavior, locomotor ability, speech, hearing, and cognition
(Chan and Kilby, 2000). Fetuses and infants are more sensitive than adults to hypothyroidism,
both because they are sensitive to small changes in thyroid hormone levels, and because the
resulting effect is more severe. Prompt supplementation of neonates with thyroid hormone can
restore neurodevelopmental function.
2.5.2. Systemic Toxicity
Several human studies examined the short-term toxicity of chlorite and chlorine dioxide
(Lubbers et al., 1981, 1982, 1984). In these studies, groups of 10 healthy adult males drank
solutions providing a dose of 0 or 0.034 mg/kg chlorite or 0.34 mg/kg chlorine dioxide (divided
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into 2 portions, separated by 4 hours); a second study involved the ingestion of distilled water
providing 0 or 0.04 mg/kg/day chlorite or chlorine dioxide for 12 weeks. Neither of these studies
found any physiologically relevant alterations in general health, vital signs, hematologic
parameters or serum T3 or T4 levels. No data were identified on chlorite or chlorine dioxide-
induced systemic toxicity in children.
A number of animal studies have examined the subchronic and chronic toxicity of chlorite
and chlorine dioxide; however, the results are not consistent and no firm conclusions can be made
concerning the systemic toxicity of these compounds. Harrington et al. (1995a) reported stomach
lesions and alterations in spleen and adrenal weights in rats administered sodium chlorite via
gavage for 13 weeks. Haag (1949) reported treatment-related renal pathology in rats exposed to
sodium chlorite in drinking water for 2 years, although the sensitivity of the study was limited by
the use of a small number of animals.
Daniel et al. (1990) reported increases in nasal lesions in rats exposed to chlorine dioxide
in drinking water for 90 days, but it is not known whether the nasal lesions resulted from inhaling
chlorine dioxide vapors at the drinking water tube or from off-gassing of the vapors. Haag (1949)
did not observe renal pathology in rats exposed to chlorine dioxide in drinking water for 2 years at
slightly higher molar concentrations than those that caused renal pathology in a similar study
(Haag, 1949) with rats exposed to chlorite.
Studies on the hematological effects of chlorite have shown mixed results. Moore and
Calabrese (1982) found an increase in mean corpuscular volume and osmotic fragility in mice
exposed to sodium chlorite in drinking water for 30 days, while Bercz et al. (1982) reported
decreases in erythrocyte and hemoglobin levels, and CMA (1996) reported decreased red blood
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cell parameters in F, rats at the high dose. Two studies reported decreases in osmotic fragility,
blood glutathione levels, and the activity of erythrocyte glutathione peroxidase and blood catalase
in rats exposed to chlorite in drinking water for up to a year (Abdel-Rahman et al., 1984a; Couri
and Abdel-Rahman, 1980). However, a consistent dose-response relationship was not observed
in these studies.
Studies of hematological effects of chlorine dioxide generally have been negative. Three
studies (Bercz et al., 1982; Daniel et al., 1990; Moore and Calabrese, 1982) found no hematologic
effects in monkeys, rats, or mice exposed to chlorine dioxide in drinking water for 4-6 weeks, 90
days, or 30 days, respectively. Abdel-Rahman et al. (1984a) exposed rats to chlorine dioxide in
drinking water for up to 11 months, and reported increased osmotic fragility and decreased
hematocrit and hemoglobin levels, but there was no clear dose response, and the effects were not
consistent across interim measurement periods. Two studies (Abdel-Rahman et al., 1984a; Couri
and Abdel-Rahman, 1980) reported decreases in erythrocyte glutathione levels, increases in
glutathione peroxidase activity, and increases in erythrocyte catalase levels in rats exposed to
chlorine dioxide in drinking water for up to one year. However, a consistent dose-response
relationship was not observed in these studies.
2.5.3. Carcinogenicity
There are no studies on the carcinogenicity of chlorite in humans. There is one animal oral
carcinogenicity study on chlorite (Kurokawa et al., 1986a). In this study, rats and mice were
exposed to sodium chlorite in drinking water for 80 or 85 weeks at doses up to 32 mg/kg/day for
male rats, 41 mg/kg/day for female rats, and 71 mg/kg/day for male and female mice. Significant
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increases in liver and lung tumors were observed in the male mice, but these incidences were
within the range of historical controls in the laboratory. The mouse study is considered
inadequate for assessing carcinogenicity because of the relatively short duration and the high
incidence of early mortality in the control male mice. In rats, the only effect was a slight (<10%)
dose-related decrease in body weight gain; there were no exposure-related increases in tumors in
rats.
There have been no studies in humans and no long-term oral bioassays on the
carcinogenicity of chlorine dioxide in animals. In a short-term assay (Miller et al., 1986), partially
hepatectomized rats received a single dose of concentrated water (chlorine dioxide concentration
not reported) in 2% emulphor followed 1 week later by administration of 500 mg/L phenobarbital
in drinking water for 56 days. No significant increases in the incidence of y-
glutamyltranspeptidase-positive foci (a measure of preneoplastic change) were observed.
Chlorite and chlorine dioxide have shown mixed results in mutagenicity assays. Both
chlorite and chlorine dioxide induced reverse mutations in S. typhimurium (with S9 activation)
(Ishidate et al., 1984). However, in a study by Miller et al.( 1986), water samples disinfected with
chlorine dioxide did not induce reverse mutations in S. typhimurium (with or without activation).
Chlorite increased the incidence of chromosomal aberrations in Chinese hamster fibroblast cells,
while chlorine dioxide did not (Ishidate et al., 1984). The results for chlorite and chlorine dioxide
from in vivo assays, such as the bone-marrow chromosomal aberration and the sperm-head
abnormality assays in mice, are primarily negative (Hayashi et al., 1988; Meier et al., 1985).
The available data do not provide sufficient evidence to support conclusions as to the
carcinogenic potential of chlorite or chlorine dioxide. Following the EPA's 1986 Guidelines for
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Carcinogen Risk Assessment, chlorite and chlorine dioxide are best classified as Group D: Not
Classifiable as to Human Carcinogenicity. This classification is appropriate because no data were
identified on human carcinogenicity and there are only preliminary animal carcinogenicity data.
Under the 1999 draft guidelines (EPA, 1999), the data for chlorine dioxide and for chlorite are
inadequate for an assessment of human carcinogenic potential.
2.5.4. Basis for RfD, MCLG, and MRDLG
EPA based the RfD for both chlorite and chlorine dioxide on the CMA (1996) two-
generation study, which identified a NOAEL of 3 mg/kg/day chlorite based on lowered auditory
startle amplitude and decreased liver weight. An UF of 100 was applied for extrapolation from an
animal study to humans and to account for variation in sensitivity among members of the human
population (EPA, 2000c). The resulting RfD is 0.03 mg/kg/day. The MCLG for chlorite and the
revised MRDLG for chlorine dioxide are both calculated to be 0.8 mg/L for a 70 kg adult
drinking 2 L of water per day. A number of studies support both the conclusion that
neurodevelopmental toxicity is the critical effect, and the identified NOAEL. The principal study
(CMA, 1996) is closely supported by the study of Mobley et al. (1990), who found a
neurobehavioral NOAEL of 3 mg/kg/day and a LOAEL of 6 mg/kg/day, based on decreased
exploratory activity in pups exposed to chlorite during gestation and lactation. Further support
for neurodevelopmental toxicity being the critical effect and for the NOAEL comes from Orme et
al. (1985), who found a chlorine dioxide NOAEL of 3 mg/kg/day and a LOAEL of 14 mg/kg/day,
based on neurobehavioral effects in rat pups exposed during gestation and lactation.
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2.5.5. Children's Risk in Relation to the MCLG and MRDLG
The MCLG and MRDLG calculated for chlorite and chlorine dioxide are considered to be
protective of sensitive subpopulations, including children, because the RfD is based on a NOAEL
derived from a two-generation reproductive toxicity study that examined numerous
developmental, reproductive and systemic endpoints. In addition, the results of this study are
supported by the results of four other developmental studies that showed similar effects at similar
dose levels. The UF of 100 used in the derivation of the RfD includes a 10-fold factor to account
for human variability in response to the toxic effects of these chemicals, including the response of
sensitive individuals such as children.
2.6. CHLORINE
Chlorine forms elemental chlorine (Cl2), chloride ion (Cl~), and hypochlorous acid (HOC1)
in pure water. As pH increases, hypochlorous acid dissociates to hypochlorite ion (OCl~).
Several factors, including chlorine concentration, pH, temperature, exposure to light, and
presence of catalysts or organic material affect the stability of free chlorine in aqueous solution.
Because hypochlorite solutions are more stable than hypochlorous acid, calcium hypochlorite and
sodium hypochlorite are often used as chlorine sources for disinfection of drinking water (EPA,
1994d). Chlorine and hypochlorites are very reactive and thus can react with the constituents of
saliva and possibly food and gastric fluid to yield a variety of reaction byproducts. Thus, the
health effects associated with administration of high levels of chlorine and/or the hypochlorites in
various animal studies may be due to these reaction byproducts and not the disinfectant itself
(EPA, 1994d).
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Scully and White (1991) noted that reactions of aqueous chlorine with sulfur-containing
amino acids appear to be so fast in saliva that all free available chlorine is dissipated before water
is swallowed (EPA, 1994d). Therefore, there is very limited potential for oral exposure of
fetuses, infants and children to chlorine.
2.6.1. Developmental/Reproductive Effects
Animal studies have demonstrated no evidence of developmental effects associated with
chlorine (EPA, 1994d). In a study by Carlton and Smith (1985), developmental landmarks such
as the mean day of eye opening and the average day of observed vaginal patency were compared
across groups with no statistical differences detected. In this study, chlorine was administered by
gavage in deionized water at doses of 1.0, 2.0, and 5.0 mg chlorine per kg/day to male and female
Long-Evans rats for 66-76 days. No statistical differences were observed between the control
and dosed groups in litter survival, litter size and pup weight. The NOAEL in this study is 5
mg/kg/day; however, higher doses were not tested (EPA, 1994h).
In a multigenerational study, rats were given drinking water chlorinated to a concentration
of 100 mg free chlorine/L (14 mg/kg/day) (Druckrey, 1968). The term "free chlorine" (free
available chlorine, free residual chlorine) refers to the concentrations of elemental chlorine,
hypochlorous acid, and hypochlorite ion that collectively occur in water. Animals were mated
repeatedly and continued to drink the test water throughout gestation and lactation.
Microphthalmia of one or both eyes was noted in 17 treated progeny but it was stated that this
condition has been known to occur spontaneously in BDII rats. No adverse reproductive or
developmental effects were observed.
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Meier et al. (1985) demonstrated that oral administration of a sodium hypochlorite
solution resulted in dose-related increases in the number of sperm-head abnormalities in male
B6C3F1 mice. Ten animals/group were given 1 mL of a free-residual-chlorine solution daily for
5 days. Test solutions were prepared by bubbling Cl2 into a 1 M solution of NaOH and adjusted
to a pH of either 8.5 (predominant species OC1") or 6.5 (predominant species HOC1). The
solutions were diluted with distilled water to 200 mg/L, 100 mg/L, and 40 mg/L chlorine
equivalents (8.0, 4.0, or 1.6 mg/kg/day, respectively). The mice were then sacrificed at 1, 3, or 5
weeks after the last dose was administered. In mice given OC1", significant increases in sperm-
head abnormalities were observed only at the 3-week interval at doses of 1.6 and 4.0 mg/kg/day.
These results were reproduced in retrials of the experiment. A similar level of increase compared
to the concurrent control value was seen in both the initial and repeat experiments, reaching about
twice the control values at 4.0 mg/kg/day. At 8.0 mg/kg/day, no further increases were seen
above that seen at 4.0 mg/kg/day. A limitation of this experiment was the wide variability in
response seen in the control animals. In addition, no dose of HOC1 was associated with increases
in sperm-head abnormalities. This result was surprising because OC1" should be converted to
HOC1 under the acidic conditions of the stomach and thus similar results for both species of
chlorine are expected. In light of the wide variability in the controls, the lack of a clear dose-
response, and the inconsistent results between the two species of chlorine, no reliable NOAEL or
LOAEL could be determined from this study.
Six Sprague-Dawley rats were administered 0, 1, 10, or 100 mg HOC1/L in drinking water
for 2.5 months prior to mating (Abdel-Rahman et al., 1982). Animals were maintained on the
treated water after pregnancy was confirmed (day 0) and killed on day 20. Maternal weight at
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time of death was not reported. Incidence of fetal anomalies associated with exposure to
hypochlorous acid solutions was not found to be statistically significant. Mean fetal weights from
the 10 and 100 mg/L groups were less than the control, but this decrease was not statistically
significant. There was also no significant difference between control and treated groups in
numbers of resorptions. Examination of general trends in the study indicated an increase (not
significant) in skeletal anomalies in animals treated with 10 mg HOC1/L. Soft tissue anomalies for
the 100 mg HOC1/L treatment group were increased significantly compared with the control. The
findings of these experiments were limited by the small number of study animals. In addition,
some calculations of anomaly percentages were reported incorrectly.
2.6.2. Systemic Toxicity
NTP (1992a) conducted a 2-year bioassay in which male and female F344 rats and
B6C3F1 mice were given chlorine in distilled drinking water at levels of 0, 70, 140, or 275 mg/L
(0, 4, 8, and 14 mg/kg/day). No effects on body weight, survival, or evidence of non-neoplastic
lesions were observed for any of the treated groups of animals. In this study, dosing began when
rats and mice were as young as 7 weeks old.
2.6.3. Carcinogenicity
The carcinogenicity of chlorine was tested in F344 rats and B6C3F1 mice by oral
exposure to chlorine in distilled drinking water as hypochlorite at levels up to 275 mg/L (14
mg/kg/day) over a 2-year period (NTP, 1992a). The incidence of mononuclear cell leukemia was
significantly increased in mid-dose, but not high-dose, female rats. There was no other evidence
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of carcinogenicity in rats or mice of either sex. NTP (1992a) concluded that there was equivocal
evidence of carcinogenic activity of chlorinated water in female F344/N rats based on an increase
in the incidence of mononuclear cell leukemia. In addition, NTP concluded that there was no
evidence of carcinogenic activity of chlorinated water in male F344/N rats or B6C3F1 mice of
either sex at the concentrations tested. There was also no evidence of systemic toxicity,
indicating that the maximum tolerated dose (MTD) may not have been reached, and making this
study inadequate for fully assessing the carcinogenic potential of chlorine. Under EPA's 1986
Guidelines for Carcinogen Risk Assessment (EPA, 1986), EPA has categorized chlorine in Group
D: Not Classifiable as to Human Carcinogenicity (EPA, 1994a). Under the draft guidelines (EPA,
1999), the data on chlorine are inadequate for an assessment of human carcinogenic potential.
2.6.4. Basis for RfD and MRDLG
EPA based the RfD for chlorine on the NOAEL of 14 mg/kg/day identified for female rats
in a 2-year bioassay (NTP, 1992a). An UF of 100 was used to account for interspecies
extrapolation and human variability, resulting in an RfD of 0.1 mg/kg/day (EPA, 1994d). This
corresponds to an MRDLG of 4 mg/L, based on a 70 kg adult consuming 2 liters of water per
day.
2.6.5. Children's Risk Relative to the MRDLG
The Agency believes that the proposed MRDLG of 4 mg/L is protective of children's
health. Since there are no reliable data to suggest chlorine-induced developmental or
reproductive toxicity, the current RfD is based on a free standing NOAEL of 14 mg/kg/day from
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a 2-year study in which chlorine dosing began when rats and mice were as young as 7 weeks old.
The UF of 100 used in the derivation of the RfD includes a 10-fold factor to account for human
variability in response to the toxic effects of these chemicals, including the response of sensitive
individuals such as children.
2.7. CHLORAMINE
Inorganic chloramines are alternative disinfectants that are rapidly formed when free
chlorine is added to water containing ammonia. Monochloramine is the principal chloramine
formed in chlorinated natural and wastewaters at neutral pH and is much more persistent in the
environment (EPA, 1994e).
2.7.1. Developmental/Reproductive Effects
In a developmental study (Abdel-Rahman et al., 1982), the authors investigated the effects
of monochloramine administered in drinking water to female Sprague-Dawley rats. Rats were
administered 0, 1, 10, or 100 mg/L monochloramine daily in drinking water for 2.5 months before
and throughout gestation. On the 20th day of gestation, animals were sacrificed for soft tissues
and skeletal examination of the progeny. Monochloramine did not produce any significant
changes in rat fetuses at any dose level; there was a slight nonsignificant increase in fetal weight in
all chloramine-treated groups compared with controls. No reproductive toxicity studies of
chloramine were located.
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2.7.2. Systemic Toxicity
A study investigated the effects of monochloramine administered in drinking water to rats
at concentrations of 0, 1, 10, or 100 mg/L (doses of 0, 0.12, 1.2, and 12 mg/kg/day) for 12
months (Abdel-Rahman et al., 1984b). Treatment-related decreases in blood glutathione levels
were observed at 6 and 12 months, but changes in glutathione levels were inconsistent at other
times. A significant increase in red blood cell fragility was detected after 2 and 10 months of
treatment at the mid-dose level, and after 2 and 6 months at the high-dose level. At 1.2 and 12
mg/kg/day, significant changes in hematologic parameters, including decreased red blood cell
count and hematocrit, were observed at 3 months, but at the 10-month evaluation the only
affected hematologic parameters were decreased hemoglobin concentration and mean corpuscular
hemoglobin at the high dose. The increased osmotic fragility observed was not corroborated by
NTP (1992a) and was not affected in the acute experimental series. The biological significance of
these changes is uncertain, in light of inconsistent findings at different exposure durations, and in
the absence of food consumption and water intake data.
NTP (1992a) exposed F344/N rats to 0, 50, 100, or 200 ppm (2.8, 5.3, and 9.5
mg/kg/day) chloramine in drinking water. At the highest dose tested (9.5 mg/kg/day) there were
statistically significant changes in body weight, but mean body weights were within 10% of
controls until week 97 for females and week 101 for males. Decreases in liver and kidney weight,
and increases in the brain- and kidney-to-body weight ratios in the high dose males were related to
lower body weights and were not considered toxicologically significant. The test animals
consumed a reduced amount of water, which was perhaps due to unpalatability, and NTP did not
consider these changes in body weight biologically significant. In a related study, NTP (1992a)
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administered chloramine to B6C3F1 mice in drinking water at 0, 50, 100, or 200 ppm. Water
consumption was decreased at the high dose, but no treatment-related effects were observed. The
NOAEL identified in this study was the high dose, 9.5 mg/kg/day.
2.7.3. Carcinogenicity
NTP (1992a) tested chloramine in drinking water in a 2-year bioassay in male and female
F344 rats and B6C3F1 mice. NTP concluded that there was no evidence of carcinogenicity in
male rats or in mice of either sex at the doses tested. In female rats, mononuclear cell leukemia
was marginally increased at the mid and high doses. Based on these increases, NTP concluded
that there was equivocal evidence of carcinogenicity in female rats. In the absence of evidence of
noncancer effects, it is unclear whether the assay reached the MTD. Under EPA's 1986
Guidelines for Carcinogen Risk Assessment (EPA, 1986), the EPA has categorized
monochloramine in Group D: Not Classifiable, because of inadequate human and animal evidence
(EPA, 1994a). Under the draft guidelines (EPA, 1999), the data on chloramine are inadequate
for an assessment of human carcinogenic potential.
2.7.4. Basis for the RfD and MRDLG
EPA based the chloramine RfD on the absence of adverse effects in the lifetime study in
rats (NTP, 1992a). The high dose of 9.5 mg/kg/day was the NOAEL. An UF of 100 was applied
(10 for interspecies differences and 10 for intraspecies variability), resulting in an RfD of 0.1
mg/kg/day. An MRDLG of 3 mg/L for chloramine (4 mg/L measured as total chlorine) was
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derived, based on the lack of toxic effects, for a 70 kg adult consuming 2 L/day of water and
assuming a relative source contribution (RSC) from drinking water of 80%.
2.7.5. Children's Risk in Relation to the MRDLG
Only limited data are available for comparing the effects of chloramine on fetuses and
children with the effects in adults. No developmental effects were observed in the single
available developmental toxicity study (Abdel-Rahman et al., 1982). Monochloramine caused
hematological effects in rats exposed to 100 mg/L for 12 months (Abdel-Rahman et al, 1984b),
Data on hematological effects of chloramine in infants and young animals are not available, but
these groups may be more sensitive than adults to these effects. People with decreased reducing
power may be more sensitive to chloramine, because such deficits make cells, particularly
erythrocytes, more vulnerable to the toxic effects of chemicals such as monochloramine. This
suggests that infants may be more sensitive than adults to hematotoxic effects of chloramine.
Newborns may also be more susceptible, because the reducing power in red blood cells is used to
reduce methemoglobin to hemoglobin, and infants have a transient deficiency of methemoglobin
reductase, the enzyme that reduces methemoglobin to hemoglobin. People deficient in the
glucose-6-phosphate dehydrogenase (G-6PD) enzyme are another potential sensitive population,
since they have a deficit in reducing power and in levels of reduced glutathione. This enzyme
deficiency is more frequent in male members of certain ethnic groups, such as Asians, Arabs,
Caucasians of Latin ancestry, African Americans, and Africans, indicating that these ethnic groups
may have a higher susceptibility to the hematological effects of chloramine. Nevertheless, the
Agency believes that the MRDLG for chloramine, which is based on a NOAEL and includes the
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standard UF of 10 for protection of sensitive populations, is protective of infants, although there
is some uncertainty in this conclusion in the absence of a quantitative analysis, in light of the
possibility that newborns may have increased sensitivity to the hematological effects of
chloramine.
2.8. MX [3-Chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone]
Recent reports have identified the presence of 3-chloro-4-(dichloromethyl)-5-hydroxy-
2(5H)-furanone (MX) and MX-related chlorohydroxyfuranone compounds in drinking water
samples disinfected with chlorine or other agents, such as chloramine or chlorine dioxide. MX is
produced as a by-product of chlorine disinfection of drinking water containing humic material and
as a byproduct of chlorine bleaching of pulpwood. MX was first identified in the spent liquors
from kraft pulp bleaching approximately 16 years ago (Holmbom et al., 1984). Based on recent
health assessments completed for the chlorohydroxyfuranones, it was determined that, of the
chlorohydroxyfuranones, MX is the most potent mutagen and has the more extensive database.
Data regarding possible developmental, reproductive or long-term chronic health effects for
chlorohydroxyfuranones other than MX are not available. Accordingly, this section will focus on
MX as a representative chlorohydroxyfuranone.
2.8.1. Developmental/Reproductive Effects
No data were located on the reproductive toxicity of MX. The developmental toxicity of
MX has been evaluated in a study conducted in Wistar rats. Huuskonen et al. (1997)
administered doses of 0, 3, 30, or 60 mg/kg/day MX via oral gavage to pregnant rats on gestation
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days 6-19. The animals were sacrificed on gestation day 20. Indications of maternal toxicity
(reduced body-weight gain, decreased absolute and relative kidney weights, and decreased water
consumption) were observed at the highest dose. The mean body weights of the fetuses were
slightly reduced (4-6%) at all doses, but the effect was not considered statistically significant. No
increases in gross, visceral or skeletal malformations were observed in the fetuses in dose groups
compared to controls, and fetal mortality was unaffected by MX exposure. The authors
concluded that MX was not teratogenic in this strain of rat.
Teramoto et al. (1998) evaluated the teratogenic properties of MX using the micromass in
vitro assay. Twelve-day old rat embryo midbrain (central nervous system (CNS)) and limb bud
(LB) cells were exposed to MX at concentrations of 0, 1, 2, 5 or 10 |ig/mL culture medium for 5
days in the presence or absence of S9 mix. MX had no effect on the number of differentiated foci
in CNS cells, and little or no effect on LB cells when the assay was conducted in the presence of
S9. Although MX was rapidly degraded in the culture medium, treatment in the absence of S9
mix caused a significant decrease in the number of differentiated foci in CNS (2, 5 and 10 |ig/mL)
and LB cells (5 and 10 |ig/mL), indicative of a teratogenic effect. The IC50 (concentration
estimated to result in 50% inhibition) for these effects was approximately 3 |ig/mL. Some
evidence for cytotoxicity was observed in the absence of S9 fraction in CNS cells (15% and 50%
decreases in survival at 5 and 10 |ig/mL, respectively), while MX was only weakly cytotoxic in
LB cells (15%) decrease in cell survival at 10 |ig/mL). The authors concluded on the basis of
these results that inhibition of differentiation was not a simple result of cytotoxicity. The overall
pattern of results was interpreted by the authors as evidence that MX is a direct-acting teratogen
(i.e., does not require bioactivation to exert teratogenic effects) and that microsomal metabolism
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(S9) minimizes the teratogenic risk. It should be noted, however, that the MX concentrations
used in this in vitro experiment were high relative to levels expected to occur in vivo following
ingestion through drinking water (approximately 1 x 10"6 |ig/mL based on a drinking water intake
of 2 L/day containing 70 ng/L MX, assuming 35% absorption and uniform distribution to a total
body volume of 40 L for an adult human) (Guyton, 1981).
2.8.2. Systemic Effects
MX has a strong local irritant effect on the gastrointestinal tract, as evidenced by
necrosis and hyperplasia of the duodenum and forestomach in both acute (Komulainen et al.,
1994) and subchronic (Vaittinen et al., 1995) rat studies. However, in the subchronic study
(Vaittinen et al., 1995), it is possible that the gastrointestinal lesions may have resulted from bolus
dosing. In the Komulainen et al. (1994) study, respiratory distress, ataxia, cyanosis, and reduced
motor activity were also noted. Indicators of altered renal and kidney function, including
increased blood-urea nitrogen, creatinine, liver weight and serum-cholesterol levels, were reported
in the Vaittinen et al. (1995) study, but were not accompanied by marked morphological or
histopathological changes.
2.8.3. Carcinogenicity
No data were identified on the carcinogenicity of MX in humans. In animals, MX is a
probable multiple-target organ carcinogen, with tumors observed in the thyroid, liver, mammary
gland, lung, adrenal, and pancreas in a 2-year oral carcinogenicity study in Wistar rats
(Komulainen et al., 1997). The liver and thyroid glands were found to be the primary target
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organs of tumorigenicity. MX promoted tumor formation in the glandular stomach of male
Wistar rats when it was administered following pretreatment with N-methyl-N-nitro-N-
nitrosoguanidine (MNNG) (Nishikawa et al., 1999).
MX is a potent mutagen in bacterial assay systems, including multiple test strains of S.
typhimurium and Escherichia coli, when tested in the absence of metabolic activation (DeMarini
et al., 1995; Meier et al., 1987). MX has also given consistently positive results in the absence of
S9 (or S9 status unknown) when tested in mammalian cells in vitro for the induction of gene
mutations, chromosome aberrations, DNA adducts, or other indicators of DNA damage
(Harrington-Brock et al., 1995; Le Curieux et al., 1999; Maki-Paakkanen et al., 1994; Meier et
al., 1987, 1989). These results have been observed, however, at concentrations significantly
higher than those expected to occur in vivo. Evidence from in vivo studies is equivocal, with
negative results obtained in assays based on micronucleus formation in bone marrow and
peripheral blood (Jansson, 1998; Meier et al. 1987, 1996) and positive results obtained for
micronuclei induction and sister-chromatid exchange in peripheral lymphocytes (Jansson et al.,
1995; Maki-Paakkanen and Jansson, 1995). Additional evidence for direct DNA damage has
been obtained in the form of strand breaks in multiple tissues and nuclear anomalies in intestinal
cells (Daniel et al., 1991; Sasaki et al., 1997). Overall, the weight of evidence indicates that MX
is a direct-acting genotoxicant in mammalian cells.
Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, MX is best
classified as Group B2: Probable Human Carcinogen. This classification is appropriate because
there is sufficient evidence in animals and inadequate evidence in humans. Under the 1999 Draft
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Guidelines for Carcinogen Risk Assessment (EPA, 1999), MX is likely to be carcinogenic to
humans.
EPA selected the 2-year oral-exposure study conducted in rats by Komulainen et al.
(1997) for quantitative evaluation of the carcinogenic effects of MX. Data for occurrence of
thyroid follicular-cell adenoma and adenocarcinoma, and liver cholangioma and
cholangiocarcinoma were modeled. Linear low-dose extrapolation was used because the weight
of evidence supports the view that MX is a strong bacterial mutagen and is also mutagenic in
mammalian cells in vitro. Using the data for thyroid follicular adenomas in male rats (as the most
sensitive tumor type) and linear extrapolation from the LED10, a cancer oral slope factor of 3.7
(mg/kg/day)"1 was calculated. Based on this oral slope factor, drinking water concentrations of
approximately 0.95 //g/L, 0.095 //g/L, and 0.0095 //g/L are associated with lifetime cancer risks
of 10"4, 10"5, and 10"6, respectively.
2.8.4. Basis for RfD and MCLG
EPA has not derived an RfD for MX. There are insufficient data available to evaluate
whether children are more sensitive to the toxic effects of MX than are adults.
2.8.5. Children '.s Risk in Relation to the MCLG
The Agency is not proposing an MCLG for 3-chloro-4-(dichloromethyl)-5-hydroxy-
2(5H)-furanone (MX) at the present time because the Agency has not yet conducted a full
assessment of the carcinogenicity of MX.
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3. SUMMARY AND CONCLUSIONS
In developing the Proposed Stage 2 D/DBP Rule, risks to sensitive subpopulations
including fetuses and children were taken into account in the assessments of D/DBPs. To
determine whether fetuses and children are more sensitive than adults, the following issues were
considered:
1. Is there information that shows that the D/DBP causes effects in the developing
fetus or harms a woman's ability to become pregnant and bear children? If it
causes these effects, do these effects occur at lower doses than those that cause
other types of effects?
2. If the D/DBP causes a health effect other than cancer, such as an effect on the liver
or kidney, are children affected at lower doses than are adults?
3. If the D/DBP causes cancer, are children more likely to be affected by a given dose
than are adults?
The ultimate goal of these questions is to determine whether the MCLG is protective of
any putative special risk to children, regardless of whether the MCLG is based on developmental
toxicity, systemic toxicity, or cancer effects.
Table ES-1 and Table 3 summarized the comparison of toxicity endpoints for the various
D/DBPs. As can be seen in the table, bromodichloromethane (BDCM), bromoform,
dichloroacetic acid (DCA), bromate, and 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone
(MX) are considered probable carcinogens for humans under the 1986 cancer guidelines and as
likely to be carcinogenic to humans under the revised 1999 draft cancer guidelines. MCLGs of
zero were selected after consideration of the potential carcinogenicity of these chemicals, except
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for MX, for which the MCLG has not yet been determined. The MCLG of zero would protect
both children and adults. The MCLG for chloroform was set based on a nonlinear extrapolation
to low doses for the cancer assessment. Liver toxicity was considered the most sensitive effect
for chloroform and as a key event in carcinogenicity. This approach is considered equally
protective for children and adults because the database does not indicate that children are more
sensitive than adults to liver toxicity. In addition, the mode of action by which chloroform
produces organ toxicity and carcinogenicity is considered to be the same in children and adults.
The MCLG/MRDLGs for dibromochloromethane (DBCM), monochloroacetic acid
(MCA), trichloroacetic acid (TCA), chlorine, and chloramine were based on systemic toxicity.
The NOAEL/LOAELs used to derive these numbers are lower than the NOAEL/LOAELs for
developmental effects; therefore, the MCLG/MRDLG would be protective of developmental
effects in infants and children. Based on the available data, it is believed that the MCLGs for
these chemicals are also protective of systemic effects in children, in light of the absence of
evidence to the contrary, and because the UFs used in the derivation of these RfDs includes a 10-
fold factor to account for human variability in response to the toxic effects of these chemicals,
including the response of sensitive individuals such as children. There is, however, some
uncertainty in this conclusion, since inadequate data are available for these chemicals to directly
compare systemic toxicity in children and adults, or to extrapolate based on data on age-related
differences in metabolic capacity. In the case of chlorine, the MRDLG is based on systemic
toxicity because the NOAEL of 5 mg/kg/day based on developmental effects is the highest dose
tested and is, therefore, not a true NOAEL. For chlorine dioxide and chlorite, the
MCLG/MRDLGs are calculated based on data from developmental studies; hence the numbers
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derived would be protective of developmental effects in both children and adults. The chlorine
dioxide/chlorite MCLG/MRDLG is also considered to be protective of systemic effects in
children, because it is based on data from a two-generation reproduction study that examined
numerous developmental, reproductive and systemic endpoints. For chloramine, the MRDLG
was based on a NOAEL of 9.5 mg/kg/day in a rat lifetime study. Infants may be more sensitive
than adults to the critical effect of chloramine (hematotoxicity). The Agency believes that the
chloramine MRDLG, which includes the standard UF of 10 for protection of sensitive
populations, is protective of infants, although there is some uncertainty in this conclusion. The
data on monobromoacetic acid (MBA), bromochloroacetic acid (BCA), dibromoacetic acid
(DBA), and 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) are insufficient for the
derivation of an MCLG. It can be concluded that the MCLG/MRDLGs of all the D/DBPs in the
proposed Stage 2 D/DBP Rule are protective of fetuses, infants and children.F
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4. REFERENCES
Abdel-Rahman M.S., M.R. Bernardi and R.J. Bull. 1982. Effect of chlorine and monochloramine
in drinking water in the developing rat fetus. J Appl Toxicol. 2(3): 156-9. (As cited in EPA
1998c)
Abdel-Rahman, M.S., D. Couri and R.J. Bull. 1984a. Toxicity of chlorine dioxide in drinking
water. J Am Coll Toxicol. 3: 277-284. (As cited in EPA 2000c)
Abdel-Rahman, M. S., Couri, D., Bull, R. J. 1984b. Toxicity of monochloramine in rat: An
alternative drinking water disinfectant. J Toxicol Envir Health. 13: 825-834. (As cited in EPA
1994e, 1992b)
Aida, Y., K. Yasuhara, K. Takada, Y. Kurokawa and M. Tobe. 1992. Chronic toxicity of
microencapsulated bromodichloromethane administered in the diet to Wistar rats. J Toxicol Sci.
17(2): 51-68, [Erratum] 17(3): 167. (As cited in EPA 2002c)
Andrews, J.E., J. Schmidt, H. Nichols, E.S. Hunter and G. Klinefelter. 1999. Developmental
toxicity of structurally related disubstituted haloacetic acids in embryo culture. Toxicol Sci. 48(1
suppl): 16. (As cited in EPA 2002b)
Anna, C., R. Maronpot, M. Pereira, J. Foley, D. Malarkey and M. Anderson. 1994. Proto-
oncogene activation in dichloroacetic acid-, trichloroethylene- and tetrachloroethylene-induced
liver tumors in B6C3F1 mice. Environmental Carcinogenesis Program, National Institute of
Environmental Health Sciences, pp. 2255-2261. (As cited in EPA 2001c)
Aschengrau, A., S. Zierler and A. Cohen. 1993. Quality of community drinking water and the
occurrence of late adverse pregnancy outcomes. Arch Environ Health. 48(2): 105-13. (As cited
in EPA 2002c)
Aschengrau, A., S. Zierler and A. Cohen. 1989. Quality of community drinking water and the
occurrence of spontaneous abortions. Arch. Environ. Health. 44: 283-90.
Austin, E.W., J.M. Parrish, D.H. Kinder and R.J. Bull. 1996. Lipid peroxidation and formation of
8-hydroxydeoxyguanosine from acute doses of halogenated acetic acids. Fundam and Appl
Toxicol. 31: 77-82. (As cited in EPA 2002b)
Bantle, J.A., R.A. Finch, D.J. Fort, E.L. Stover, M. Hull, M Kumsher-King and A.M. Gaudet-
Hull. 1999. Phase III Interlaboratory Study of FETAX Part 3. FETAX validation using 12
compounds with and without an exogenous metabolic activation system. J Appl Toxicol. 19:
447-472. (As cited in EPA 2001c)
181
-------
Barton, H. A., R. Bull, I. Schultz andM.E. Anderson. 1999. Dichloroacetate (DCA) dosimetry:
interpreting DCA-induced liver cancer dose response and the potential for DCA to contribute to
trichloroethylene-induced liver cancer. Toxicol Lett. 106: 9-21. (As cited in EPA 2001c)
Bielmeier, S.R., D.S. Best, D.L. Guidici and M.G. Narotsky. 2001. Pregnancy loss in the rat
caused by bromodichloromethane. Toxicol Sci. 59(2): 309-315. (As cited in EPA 2002c)
Benson, C. I. 1951. Potassium bromate poisoning. Br Med J. 1:1516. (As cited in EPA
2001b)
Bentley, P., Calder, I., Elcombe, C., Grasso, P., Stringer, D., and Wiegan, H.J. 1993. Hepatic
peroxisome proliferation in rodents and its significance for humans. Food Chem. Toxicol. 31:
857-907. (As cited in EPA 2002a)
Bercz, J. P., L.L. Jones, Garner, L. et al. 1982. Subchronic toxicity of chlorine dioxide and
related compounds in drinking water in the nonhuman primate. Environ Health Perspect. 46: 47-
55. (As cited in EPA 2000c)
Borzelleca J. F. and R.A. Carchman, 1982. Effects of selected organic drinking water
contaminants on male reproduction. Research Triangle Park, NC: U.S. Environmental
Protection Agency. EPA 600/1-82-009. NTIS PB82-259847. (As cited in EPA 1998c)
Boutelet-Bochan H., Y. Huang and M.R. Juchau. 1997. Expression of CYP2E1 during
embryogenesis and fetogenesis in human cephalic tissues: implications for the fetal alcohol
syndrome. Biochem Biophys Res Commun. 238: 443-7. (As cited in EPA 1998b)
Bove, F., Y. Shim, and P. Zeitz. 2002. Drinking water contaminants and adverse pregnancy
outcomes: A review. Environ Health Perspect. 110:61-74
Bove, F.J., M.C. Fulcomer, J.B.Klotz, J. Esmart, E.M. Dufficy and J.E. Savrin. 1995. Public
drinking water contamination and birth outcomes. Am J Epidemiol. 143(11): 1179-80. (As
cited in EPA 1998c)
Bove, F.J., M.C. Fulcomer, J.B. Koltz, J. Esmart, E.M. Dufficy and R.T. Zagraniski. 1992.
Report on phase IV-A: Public drinking water contamination and birthweight fetal deaths, and
birth defects, a cross-sectional study. New Jersey Department of Health. (As cited in Reif et al.
2000; WHO 2000)
Brenniman, G.R., J. Vasilomanolakis-Lagos, J. Amsel, T. Namekata and A.H. Wolff. 1980.
Case-control study of cancer deaths in Illinois communities served by chlorinated or
nonchlorinated water. In: Jolley, R. et al., eds. Water chlorination. Environmental impact and
health effects, Vol. 3. Ann Arbor, MI: Ann Arbor Science Publishers, pp. 1043-1057. (As cited
in EPA 2002c)
182
-------
Bryant, B.J., Jokinen, MP., Eustis, S.L., Thompson, M.B., Abdo, K.M. Toxicity of
monochloroacetic acid administered by gavage to F344 rats and B6C3F1 mice for up to 13
weeks. Toxicology. 72(1): 77-87, 1992. (As cited in EPA 2002a)
Bull, R.J., J.M. Brown, E.A. Henry, T.A. Jorgenson, M. Robinson and J.A. Stober. 1986.
Enhancement of the hepatotoxicity of chloroform in B6C3F1 mice by corn oil: implications for
chloroform carcinogenesis. Environ Health Perspect. 69: 49-58. (As cited in EPA 1998c)
Bull, R.J., I.M. Sanchez, M.A. Nelson, J.L. Larson and A.J. Lansing. 1990. Liver tumor induction
in B6C3F1 mice by dichloroacetate and trichloroacetate. Toxicology. 63: 341-59. (As cited in
EPA 1998c)
Bull, R.J. 2000. Mode of action of liver introduction by trichloroethylene and its metabolites,
trichloroacetate and dichloroacetate. Environ Health Perspect. 108(Suppl. 2): 241-259. (As
cited in EPA 2002a)
Cantor, K.P., C.F. Lynch, M.E. Hildesheim, M. Dosemeci, J. Lubin, M. Alavanja and G. Craun.
1998. Drinking water source and chlorination byproducts. I. Risk of bladder cancer.
Epidemiology. 9(1): 21-8. (As cited in EPA 200le)
Cantor, K.P., C.F. Lynch, M.E. Hildesheim et al. 1999. Drinking water source and chlorination
byproducts in Iowa. III. Risk of brain cancer. Amer Journal Epidemiology. 150(6): 552-560. (As
cited in EPA 200le)
Carlton, B.D. and M.K. Smith. 1985. Reproductive effects of alternate disinfectants and their by-
products. In: Jolley, R.L., et al., eds. Water chlorination: environmental impact and health
effects, vol. 5. Chelsea, MI: Lewis Publications, pp. 295-305. (As cited in EPA 1994d, 2000c)
Carlton, B.D., D.L. Habash, A.H Barsaran, E.L. George and M.K. Smith. 1987. Sodium chlorite
administration in Long-Evans rats: reproductive and endocrine effects. Environ Res. 42: 238-
245. (As cited in EPA 2000c)
Carlton, B.D., A.H. Basaran, L.E. Mezza, E.L. George and M.K. Smith. 1991. Reproductive
effects in Long-Evans rats exposed to chlorine dioxide. Environ Res. 56: 170-177. (As cited in
EPA 2000c)
Carpenter, S.P., Lasker, J.M., Raucy, J.L. 1996. Expression, induction, and catalytic activity of
the ethanol-inducible cytochrome P450 (CYP2E1) in human fetal liver and hepatocytes. Mol
Pharmacol. 49: 260-8. (As cited in EPA 1998b)
CCC. 2000a. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories,
Raymond G. York, Study Director). Oral (Drinking Water) Range-finding Developmental
Toxicity Study of Bromodichloromethane in Rats. Protocol No. 2403-001P. Arlington, VA:
Chlorine Chemistry Council. (As cited in EPA 2002c)
183
-------
CCC. 2000b. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories,
Raymond G. York, Study Director). Oral (Drinking Water) Developmental Toxicity Study of
Bromodichloromethane in Rats. Protocol No. 2403-003. Arlington, VA: Chlorine Chemistry
Council. (As cited in EPA 2002c)
CCC. 2000c. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories, Raymond
G. York, Study Director). Oral (Drinking water) Developmental Toxicity Study of
Bromodichloromethane in Rabbits. Protocol No. 2403-002. Arlington, VA: Chlorine Chemistry
Council. (As cited in EPA 2002c)
CCC. 2001. Chlorine Chemistry Council. Oral (Drinking Water) Two-Generation Reproductive
Study of Dibromoacetic Acid (DBA) in Rats. R.G. York, Study Director, Argus Research
Laboratory, Horsham, PA 19044. Study Protocol No. 2403-005, 1546 pages. (As cited in EPA,
2002b)
Chan, S. and M.D. Kilby. 2000. Thyroid hormone and central nervous system development. J.
Endocrin. 165: 1-8.
Chang, L.W., F.B. Daniel and A.B. DeAngelo. 1991. Analysis of DNA strand breaks induced in
rodent liver in vivo, hepatocytes in primary culture, and a human cell line by chloroacetic acids
and chloroacetaldehydes. Environ Molec Mutagen. 20:277-288. (As cited in EPA 2002a)
Chipman, J.K., J.E. Davies, J.L. Parsons, J. Nair, G. O'Neill, and J.K. Fawell. 1998. DNA
oxidation by potassium bromate; a direct mechanism or linked to lipid peroxidation? Toxicology.
126(2): 93-102 (As cited in EPA, 2001b)
Christian, M.S., R.G. York, A.M. Hoberman, L.C. Fisher, and W.R. Brown. 2002. Oral
(drinking water) two-generation reproductive toxicity study of bromodichloromethane (BDCM)
in rats. Int J Toxicol. 21(2): 115-146. (As cited in EPA, 2002c)
Christian, M.S., R.G. York, A.M. Hoberman, R.M. Diener, and L.C. Fisher. 2001a. Oral
(drinking water) developmental toxicity studies of bromodichloromethane (BDCM) in rats and
rabbits. Int J Toxicol. 20: 225-237. (As cited in EPA, 2002c)
Christian, M.S., R.,G. York, A.M. Hoberman, R.M. Diener, L.C. Fischer, and G.A. Gates.
2001b. Biodisposition of dibromoacetic acid (DBA) and bromodichloromethane (BDCM)
administered to rats and rabbits in drinking water during range-finding reproduction and
developmental toxicity studies. Int J Toxicol. 20: 239-253. (As cited in EPA 2002b)
Cicmanec, J. L., Condie, L. W., Olson, G. R., Wang, S. R. 1991. 90-day toxicity study of
dichloroacetate in dogs. Fundam Appl Toxicol. 17(2): 376-89. (As cited in EPA, 2001c)
184
-------
CMA. 1996. Chemical Manufacturers Association. Sodium chlorite: drinking water rat two-
generation reproductive toxicity study. Quintiles Report Ref. CMA/17/96. (As cited in EPA
1998c)
Cornett, R., M. James, G. Henderson, J. Cheung, A. Shroads and P. Stacpoole. 1999. Inhibition
of glutathione S-transferase and tyrosine metabolism by dichloroacetate: A potential unifying
mechanism for its altered biotransformation and toxicity. Biochemical and Biophysical Research
Communications. 262: 752-756. (As cited in EPA, 2001c)
Cosby, N. C. and W. R. Dukelow. 1992. Toxicology of maternally ingested trichloroethylene
(TCE) on embryonal and fetal development in mice and of TCE metabolites on in vitro
fertilization. Fundam Appl Toxicol. 19(2): 268-74. (As cited in EPA 2002a)
Couri, D. and M.S. Abdel-Rahman. (1980) Effect of chlorine dioxide and metabolites on
glutathione dependent system in rat, mouse and chicken blood. J Environ Pathol Toxicol. 3: 451-
460. (As cited in EPA 2000c)
Couri, D., C.H. Miller, R.J. Bull, J.M. Delphia and E.M. Ammar. 1982. Assessment of maternal
toxicity, embryotoxicity and teratogenic potential of sodium chlorite in Sprague-Dawley rats.
Environ Health Perspect. 46: 25-29. (As cited in EPA 2000c)
Cummings, A.M. and J.M. Hedge. 1998. Dibromoacetic acid does not adversely affect early
pregnancy in rats. Reproductive Toxicology. 12(4): 445-448. (As cited in EPA 2002b)
Daniel, F.B., L.W. Condie, M. Robinson et al. 1990. Comparative subchronic toxicity studies of
three disinfectants. J Am Water Works Assoc. 82: 61-69. (As cited in EPA 2000c)
Daniel, F.B., G. R. Olson, and J.A. Stober. 1991. The induction of gastrointestinal tract nuclear
anomalies in B6C3F1 mice by 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5//)-furanone and 3,4-
(dichloro)-5-hydroxy-2(5//)-furanone, mutagenic by-products of chlorine disinfection. Environ
Mol Mutagen. 17: 32-39. (As cited in EPA 2000e)
Daniel, F.B., A.B. DeAngelo, J.A. Stober, G.R. Olson, andN.P. Page. 1992.
Hepatocarcinogenicity of chloral hydrate, 2-chloroacetaldehyde, and dichloroacetic acid in male
B6C3F1 mouse. Fundam Appl Toxicol. 19(2): 159-68. (As cited in EPA 1998c)
Dawson, D.A., Bantle, J.A. 1987. Development of a reconstituted water medium and
preliminary validation of the frog embryo teratogenesis assay-Xenopus (FETAX). J. Appl.
Toxicol. 1987 Aug; 7(4)237-44. (As cited in EPA, 2002b)
DeAngelo, A.B. D.R. Geter, D.W. Rosenberg, C.K. Crary, and M.H. George. 2002. The
induction of aberrant crypt foci (ACF) in the colons of rats by trihalomethanes administered in the
drinking water. Cancer Lett. 187(l-2):25-31. (As cited in EPA, 2002c)
185
-------
DeAngelo, A.B., M.H. George and D.E. House. 1999. Hepatocarcinogenicity in the male
B6C3F1 mouse following a lifetime exposure to dichloroacetic acid in the drinking water: dose-
response determination and modes of action. J Toxicol Environ Health. 58(8): 485-507. (As
cited in EPA, 2001c)
DeAngelo, A.B., M.H George, S.R. Kilburn, T.M. Moore, and D.C. Wolf. 1998. Carcinogenicity
of potassium bromate administered in the drinking water to male B6C3F1 mice and F344/N rats.
Toxicol Pathol. 26(5): 587-94. (As cited in EPA, 1998c)
DeAngelo, A.B., F.B. Daniel, B.M. Most, and G.R.Olson. 1997. Failure of monochloroacetic
acid and trichloroacetic acid administered in the drinking water to produce liver cancer in male
F344/N rats. J Toxicol Environ Health. 52:425-445. (As cited in EPA 2002a)
DeAngelo, A.B., F.B. Daniel, B.M. Most, and G.R. Olsen. 1996. The carcinogenicity of
dichloroacetic acid in the male Fischer 344 rat. Toxicology. 114: 207-221. (As cited in EPA
2001c)
DeAngelo, A.B., F.B. Daniel, J.A. Stober and G.R. Olson. 1991. The carcinogenicity of
dichloroacetic acid in male B6C3F1 mouse. Fundam Appl Toxicol. 16: 337-47. (As cited in
EPA 1998c)
Dees, C. and C. Travis. 1994. Trichloroacetate stimulation of liver DNA synthesis in male and
female mice. Toxicology Letters. 70: 343-355. (As cited in EPA 2002a)
DeMarini, D.M., E. Perry and M.L. Sheldon. 1994. Dichloroacetic acid and related compounds:
induction of prophage in E. coli and mutagenicity and mutation spectra in Salmonella TA 100.
Mutagenesis. 9: 429-437. (As cited in EPA 2002a, 2001c)
DeMarini, D.M., A. Abu-Shakra, C.F. Felton, K.S. Patterson, and M.L. Shelton. 1995. Mutation
spectra in Salmonella of chlorinated, chloraminated or ozonated drinking water extracts:
comparison to MX. Environ Mol Mutagen. 26: 270-285. (As cited in EPA 1994e)
DeMarini, D.M., M.L. Shelton, S.H. Warren, T.M. Ross, J.Y. Shim, A.M. Richard and R.A.
Pegram. 1997. Glutathione S-transferase-mediated induction of GC->AT transitions by
halomethanes in Salmonella. Environ Mol Mutagen. 30(4): 440-447. (As cited in EPA 2002c)
Dodds, L., W. King, C. Woolcott and J. Pole. 1999. Trihalomethanes in public water supplies and
adverse birth outcomes. Epidemiology. 10(3): 233-237. (As cited in EPA 2002c)
Dodds, L. and W. King. 2001. Relation between trihalomethane compounds and birth defects.
Occup. Environ. Med. 58: 443-446.(As cited in EPA 2002c)
Druckrey, H. 1968. Chlorinated drinking water, toxicity tests, involving seven generation of rats.
Food and Cosmetics Toxicology, Vol. 6. Pergamon Press, pp. 147-54. (As cited in EPA 1994d)
186
-------
EPA. 2002a. Environmental Protection Agency. Drinking water criteria document for
monochloroacetic acid and trichloroacetic acid. Draft. June, 2002. Prepared for Health and
Ecological Criteria Division, Office of Science and Technology, Office of Water by Toxicology
Excellence for Risk Assessment (TERA) and GRAM, Inc.
EPA. 2002b. Environmental Protection Agency. Drinking water criteria document for
brominated acetic acids. Draft. June, 2002. Prepared for Health and Ecological Criteria
Division, Office of Science and Technology, Office of Water by Toxicology Excellence for Risk
Assessment {TERA) and GRAM, Inc.
EPA. 2002c. Environmental Protection Agency. Drinking water criteria document on
brominated trihalomethanes. External review draft. March 30, 2002. Prepared for Health and
Ecological Criteria Division, Office of Science and Technology, Office of Water by ICF
Consulting Group, Syracuse Research Corporation, and The Cadmus Group.
EPA. 2001a. Environmental Protection Agency. Toxicological review of chloroform. In
support of Integrated Risk Information System (IRIS). Washington, DC. Available online at
http ://www. epa.gov/iris/
EPA. 2001b. Environmental Protection Agency. Toxicological review of bromate. In support
of Integrated Risk Information System (IRIS). Washington, DC. Available online at
http ://www. epa.gov/iris/
EPA. 2001c. Environmental Protection Agency. Toxicological review of dichloroacetic acid. In
support of summary information on Integrated Risk Information System (IRIS). National Center
for Environmental Assessment, Washington, DC. March 2001. Draft.
EPA. 200Id. Environmental Protection Agency. Relative source contribution for chloroform.
Office of Water, Office of Science and Technology, Washington, DC. EPA-822-R-01-006.
EPA. 200le. Environmental Protection Agency. Epidemiologic literature on cancer and
exposure to chlorinated drinking water. Prepared for Office of Water by Research Triangle
Institute and Sciences International, Contract No. 68-C-98-195.
EPA. 200 If. Environmental Protection Agency. Peer Review of Toxicological Review of
Dichloroacetic Acid: In Support of Summary Information on the Integrated Risk Information
System (IRIS) Contract 68-C-98-189. Prepared by Versar, Inc., under Work Assignment 2-37.
September.
EPA. 2001g. Notice of opportunity to provide additional information and comment. Federal
Register 66:59593-59594
187
-------
EPA. 2000a. Environmental Protection Agency. Child-specific exposure factors handbook.
Prepared for U.S. EPA National Center for Environmental Assessment, Office of Research and
Development, by Versar, Inc.
EPA. 2000b. Environmental Protection Agency. Estimated per capita water ingestion in the
United States. April, 2000. U.S. EPA Office of Water. EPA-822-00-008.
EPA. 2000c. Environmental Protection Agency. Toxicological review of chlorine dioxide and
chlorite. In support of Integrated Risk Information System (IRIS). Washington, DC, September,
2000. Available online at http://www.epa.gov/iris/
EPA. 2000d. Environmental Protection Agency. Cytochrome P450. Prepared for Health and
Ecological Criteria Division, Office of Science and Technology, Office of Water by IS SI
Consulting Group.
EPA. 2000e. Environmental Protection Agency. Quantitative Cancer Assessment for MX.
August 11, 2000. Prepared for Health and Ecological Criteria Division, Office of Science and
Technology, Office of Water by IS SI Consulting Group, Inc.
EPA. 2000f. Environmental Protection Agency. Methodology for Deriving Ambient Water
Quality Criteria for the Protection of Human Health. Office of Water, Office of Science and
Technology. EPA-822-B-00-004.
EPA. 1999. Environmental Protection Agency. Guidelines for carcinogen risk assessment.
Office of Research and Development, Washington, DC. NCEA-F-0644. July. SAB Review draft.
EPA. 1998a. Environmental Protection Agency. Mar 13 National Primary Drinking Water
Regulations; Disinfectants and Disinfection Byproducts; Proposed Rule. Fed Reg. 63(61):
15674-15692.
EPA. 1998b. Environmental Protection Agency,
drinking water disinfection byproduct chloroform.
Water, November 1998.
Health risk assessment/characterization of the
Office of Science and Technology, Office of
EPA. 1998c. Environmental Protection Agency. Health risks to fetuses, infants, and children
(Final Stage 1 D/DBP Rule). October 29, 1998. Contributors: A. Bathija, N. Chiu, O. Conerly,
V. Dellarco, K. Khanna, M. Manibusan, Y. Patel, and R. Schoeny. Submitted by A. Bathija,
Health and Ecological Criteria Division, Office of Science and Technology, Office of Water.
EPA. 1998d. Environmental Protection Agency. U.S. Environmental Protection Agency.
Quantification of toxicological effects for brominated trihalomethanes. By ICF Inc., under
contract to The Cadmus Group, Inc. Contract No. 68-C7-0002, Subcontract No. 0002
188
-------
EPA. 1998e. Environmental Protection Agency. Dichloroacetic acid: Carcinogenicity
Identification Characterization Summary. NCEA-W-0372. (As cited in EPA 2001c)
EPA. 1997. U.S. Environmental Protection Agency. Volume I - General Factors, Exposure
Factors Handbook, Update to Exposure Factors Handbook. EPA/600/P-95/002Fa. Office of
Research and Development, National Center for Environmental Assessment, U.S. Environmental
Protection Agency, Washington, DC.
EPA. 1996a. Environmental Protection Agency. Proposed guidelines for carcinogen risk
assessment. Fed Reg. 61(79): 17959-18011.
EPA. 1996b. Environmental Protection Agency. Comparison of the effects of chemicals with
combined perinatal and adult exposure vs. adult only exposure in carcinogenesis bioassays. Office
of Pesticide Programs. http://www.epa.gov/pesticides/SAP/1996/index.htm.
EPA. 1996c. Environmental Protection Agency. Reproductive Toxicity Risk Assessment
Guidelines. Fed Reg. 61(212): 56274-56322.
EPA. 1996d. Environmental Protection Agency. Dichloroacetic acid. Available online at
http://www.epa.gov/iris/subst/0654.htm. (As cited in EPA 2001c)
EPA. 1994a. Jul 29. Environmental Protection Agency. National Primary Drinking Water
Regulations; Disinfectants and Disinfection Byproducts; Proposed Rule. Fed Reg. 59(145):
38668-38829.
EPA. 1994b. Environmental Protection Agency. Final draft for the drinking water criteria
document on trihalomethanes. Contract No. EPA/68-C2-0139. Washington, DC: U. S.
Environmental Protection Agency Health and Ecological Division, Office of Science and
Technology, Office of Water.
EPA. 1994c. Environmental Protection Agency. Draft drinking water health criteria document
for chlorine dioxide, chlorite and chlorate. Office of Science and Technology, Office of Water,
March 31, 1994.
EPA. 1994d. Environmental Protection Agency. Draft drinking water criteria document for
chlorine, hypochlorous acid and hypochlorite ion. Office of Science and Technology, Office of
Water, January 1994.
EPA. 1994e. Environmental Protection Agency. Draft drinking water health criteria document
for chloramine. March 1994.
EPA. 1994f. Environmental Protection Agency. Methods for Derivation of Inhalation Reference
Concentrations and Application of Inhalation Dosimetry, EPA/600/8-90/066F, dated October
1994.
189
-------
EPA. 1994g. Environmental Protection Agency. Final Draft for the Drinking Water Criteria
Document on Chlorinated Acids/Aldehydes/Ketones/Alcohols. Prepared for Health and
Ecological Criteria Division Office of Science and Technology, Office of Water, U.S.
Environmental Protection Agency, Washington, DC 20460. EPA 68-C2-0139.
EPA. 1994h. Environmental Protection Agency. Integrated Risk Information System. Chlorine.
Washington, DC: US Environmental Protection Agency. Available online at
http ://www. epa. gov/iris/
EPA. 1991a, Jan 30. Environmental Protection Agency. National Primary Drinking Water
Regulations. Fed Reg. 56(20): 3526-3597.
EPA. 1991b, December. Environmental Protection Agency. Guidelines for developmental
toxicity risk assessment. Federal Register 56(234) 63798-63826
EPA. 1988. Environmental Protection Agency. Recommendations for and documentation of
biological values for use in risk assessment. EPA 600/6-87-008. Cincinnati, OH: Office of Health
and Environmental Assessment, for the Office of Solid Waste and Emergency Response.
EPA. 1986. Environmental Protection Agency. Guidelines for carcinogen risk assessment. Fed
Reg. 51(185): 33992-34003.
Epstein, D.L., G.A. Nolen, J.L. Randall, S.A. Christ, E.J. Read, J.A. Stober and M.K. Smith.
1992. Cardiopathic effects of dichloroacetate in the fetal Long-Evans rat. Teratology.
46(3): 225-235. (As cited in EPA, 1998c)
Ferreira-Gonzalez, A., A.B. DeAngelo, S. Nasim and C.T. Garrett. 1995. Ras oncogene
activation during hepatocarcinogenesis in B6C3F1 male mice by dichloroacetic and trichloroacetic
acids. Carcinogenesis. 16(3): 495-500. (As cited in EPA 2001c, 2002a)
Fort, D., E. Stover, J. Rayburn, M. Hull and J. Bantle. 1993. Evaluation of the developmental
toxicity of trichloroethylene and detoxification metabolites using Xenopus. Teratogenesis,
Carcinogenesis, and Mutagenesis. 13: 35-45. (As cited in EPA 2002a)
Fu, L., E.M. Johnson and L.M. Newman. 1990. Prediction of the developmental toxicity hazard
potential of halogenated drinking water disinfection by-products tested by the in vitro hydra
assay. Regulatory Toxicology and Pharmacology. 11: 213-219. (As cited in EPA 2002a)
Fujie, K., T. Aoki and M. Wada. 1990. Acute and subacute cytogenetic effects of the
trihalomethanes on rat bone marrow cells in vivo. Mutat Res. 242(2): 111-119. (As cited in
EPA 2002c)
190
-------
Fuscoe, J.C., A.J. Afshari, M.H. George, A.B. DeAngelo, R.R. Tice, T. Salman and J.W. Allen.
1996. In vivo genotoxicity of dichloroacetic acid: evaluation with the mouse peripheral blood
micronucleus assay and the single cell gel assay. EnvMolMutag. 27:1-9. (As cited in EPA
2001c)
Gallagher, M.D., J.R. Nuckols, L. Stallones and D.A. Savitz. 1998. Exposure to trihalomethanes
and adverse pregnancy outcomes. Epidemiology. 9(5): 484-489.
Galloway, S.M., M. J. Armstrong, C. Reuben, S. Colman, B. Brown, C. Cannon, A.D. Bloom, F.
Nakamura, M. Ahmed, S. Duk, J. Rimpo, B.H. Margolin, M.A. Resnick, B. Anderson and E.
Zeiger. 1987. Chromosome aberrations and sister chromatid exchanges in Chinese hamster ovary
cells: Evaluations of 108 chemicals. Environ Molec Mutagen. 10 (Suppl.): 1-175. (As cited in
EPA 2002a)
Gao, P. and R.A. Pegram. 1992. In vitro hepatic microsomal lipid and protein binding by
metabolically activated bromodichloromethane. Toxicologist 12:421. (As cited in EPA 2002c)
Gardner, H.S. and M.W. Toussant. USACEHR drinking water disinfection byproduct testing
with FETAX: Bromodichloromethane, dibromoacetic acid, and chlorinated surface water. U.S.
Army Center for Environmental Health Research. October 15, 1999. (As cited in EPA, 2002b)
George, M.H., G.R. Olson, D. Doerfler, T. Moore, S. Kilburn, and A.B. DeAngelo. 2002.
Carcinogenicity of bromodichloromethane administered in drinking water to male F344/N rats and
B6C3F1 mice. Int J Toxicol. 21(3):219-30. (As cited in EPA, 2002c)
Gerba, C.P., J.B. Rose and C.N. Haas. 1996. Sensitive populations: who is at the greatest risk?
Intl J Food Microbiol. 30:113-23. (As cited in EPA, 1998c)
Giller, S., F. Le Curieux, F. Erb and D. Marzin. 1997. Comparative genotoxicity of halogenated
acetic acids found in drinking water. Mutagenesis. 12(5): 321-328. (As cited in EPA 2002a,
2002b)
Ginsberg, G., D. Hattis, B. Sonawane, A. Russ, P. Banati, M. Kozlak, S. Smolenski, and R.
Goble. 2002. Evaluation of child/adult pharmacokinetic differences from a database derived from
the therapeutic drug literature. Toxicol Sci. 66(2): 185-200
Gradus (Ben-Ezer), D., M. Rhoads, L. B. Bergstrom and S. C. Jordan. 1984. Acute bromate
poisoning associated with renal failure and deafness presenting as hemolytic uremic syndrome.
Am J Nephrol. 4:188-191. (As cited in EPA 2000b)
191
-------
Guyton, A.C. 1981. Textbook of medical physiology. Philadelphia: W.B. Saunders Co. p. 391.
(As cited in EPA 1994e)
Haag, H.B. (1949) The effect on rats of chronic administration of sodium chlorite and chlorine
dioxide in the drinking water. Report to the Mathieson Alkali Works from H.B. Haag of the
Medical College of Virginia. February 7, 1949. (As cited in EPA 2000c)
Hakkola, J., O. Pelkonen, M. Pasanen and H. Raunio. 1998a. Xenobiotic-metabolizing
cytochrome P450 enzymes in the human feto-placental unit: role in intrauterine toxicity. Crit Rev
Toxicol. 28(1): 35-72. (As cited in EPA 1998b)
Hakkola, J., E. Tanaka, and O. Pelkonen. 1998b. Developmental expression of cytochrome P-
450 enzymes in human liver. Pharmacol Toxicol. 82(5): 209-217. (As cited in EPA 1998b,
2002c)
Hamada, F., S. Ohzono, S. Yamada, Y. Baba, Y. Tokuda, W. Yamashita, A. Nakashima, R.
Herada, T. Arima and T. Morita. 1990. A case of acute potassium bromate intoxication.
Fukuoka Igaku Zasshi. 81: 271-276. (As cited in EPA 2001b)
Harrington, R.M., R.R. Romano, Gates, D. et al. 1995a. Subchronic toxicity of sodium chlorite
in the rat. J Am Coll Toxicol. 14: 21-33. (As cited in EPA 2000c)
Harrington, R.M., Romano, R.R., Irvine, L. 1995b. Developmental toxicity of sodium chlorite in
the rabbit. J Am Coll Toxicol. 14: 109-118. (As cited in EPA 2000c)
Harrington-Brock, K., C.L. Doerr and M.M. Moore. 1995. Mutagenicity and clastogenicity of
3-chloro-4-(dichloromethyl)-5-hydroxy-2-[5H] furanone (MX) in L5178Y/TK +/(-)-3.7.2C
mouse lymphoma cells. Mutat Res. 348:105-110. (As cited in EPA 2000e)
Harrington-Brock, K. C.L. Doerr and M.M. Moore. 1998. Mutagenicity of three disinfection by-
products; di- and trichloroacetic acid and chloral hydrate in L5178Y/TK+/" - 3.7.2C mouse
lymphoma cells. Mutation Research. 413: 265-276. (As cited in EPA 2001c, 2002a)
Hayashi, M., M. Kishi, T. Sofuni and M. Ishidate Jr. 1988. Micronucleus tests in mice on 39 food
additives and eight miscellaneous chemicals. Food Chem Toxicol. 26(6): 487-500. (As cited in
EPA 2000c, 2002c)
Herren-Freund, S.L., M.A. Pereira, M.D, Khoury and G. Olson. 1987. The carcinogenicity of
trichloroethylene and its metabolites, trichloroacetic acid and dichloroacetic acid, in mouse liver.
Toxicol Appl Pharmacol. 90:183-189. (As cited in EPA 2001c)
192
-------
Heywood, R., R.J. Sortwell, P.R.B. Noel, A.E. Street, D.E. Prentice, F.J.C. Roe, P.F.
Wadsworth, A.N. Worden and N.J. Van Abbe, 1979. Safety evaluation of toothpaste containing
chloroform: III. Long-term study in beagle dogs. J Environ Pathol Toxicol. 2: 835-51. (As cited
in EPA 2001a)
Hildesheim, M.E., K.P. Cantor, C.F. Lynch, M. Dosemeci, J. Lubin, M. Alavanja and G. Craun.
1998. Drinking water source and chlorination byproducts. II. Risk of colon and rectal cancers.
Epidemiology. 9(1): 29-35. (As cited in EPA 200le)
Holmbom, B., R.H. Voss, R.D. Mortimer, and A. Wong. 1984. Fractionation, isolation and
characterization of Ames mutagenic compounds in kraft chlorination effluents. Environ Sci
Technol. 18: 333-337. (As cited in EPA 2000e)
Hunter, E.S., E.H. Rogers, J.E. Schmid and A. Richard. 1996. Comparative effects of haloacetic
acids in whole embryo culture. Teratology. 54: 57-64. (As cited in EPA 1998c)
Huuskonen, H., R. VenSalSainen, H. Komulainen, and T. Vartiainen. 1997. Developmental
toxicity of 3-chloro-4-(dichloromethyl)-5-hydroxy-2-(5H) furanone (MX) in rats. Hum Exp
Toxicol. 16(10): 61. (Abstract only). (As cited in EPA 2000e)
IARC. 2002. International Agency for Research on Cancer. Some Drinking-water disinfectants
and contaminants, including Arsenic. In: IARC Monographs on the Evaluation of Carcinogenic
Risks to Humans. 84:15-22.
IARC. 1995. International Agency for Research on Cancer. Dry cleaning, some chlorinated
solvents, and other industrial chemicals. In: IARC Monographs on the Evaluation of
Carcinogenic Risks to Humans. 63: 271-314. (As cited in EPA 2001c)
Infante-Rivard C., E. Olson, L. Jacques, and P. Ayotte. 2001. Drinking Water Contaminants and
Childhood Leukemia. Epidemiology. 12(1): 13-19. (As cited in EPA 2001 e)
ILSI. 1997. International Life Sciences Institute. An evaluation of EPA's proposed guidelines
for carcinogenic risk assessment using chloroform and dichloroacetate as case studies: Report of
an expert panel. Washington, DC. November. (As cited in EPA 2001a, 2001c)
IPCS. 2001. International Program on Chemical Safety (IPCS). Guidance Document for the Use
of Data in Development of Chemical-Specific Adjustment Factors (CSAF) for Inter-species
Differences and Human Variability in Dose/Concentration Response Assessment.
WHO/PCS/01.4. http://www.ipcsharmonize.org/csaf-intro.html#download
193
-------
Ishidate, M., T. Sofuni, K. Yoshikawa and M. Hayashi. 1982. Studies on the mutagenicity of
low boiling organohalogen compounds. Unpublished intraagency report to the National Institute
of Hygienic Sciences. Tokyo, Japan: Tokyo Medical and Dental University. (As cited in EPA
2002c)
Ishidate, M., T. Sofuni, K. Yoshikawa, et al. 1984. Primary mutagenicity screening of food
additives currently used in Japan. Food Chem Toxicol. 22: 623-636. (As cited in EPA 2000c)
Jaakkola, J.J.K., P. Magnus, A. Skrondal, J. Alexander, G. Becher, T. Krogh and E. Dybing
1999. Water chlorination and birth defects. Epidemiology. 10: S56. Abstract. (As cited in Reif
et al. 2000)
Jansson, K. 1998. Lack of induction of micronuclei in bone marrow erythrocytes of rats exposed
to 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) for 2 years in a carcinogenicity
bioassay. Environ Mol Mutagen. 32:185-187. (As cited in EPA 2000e)
Jansson, K., J.M. Hyttinen, M. Niittykoshi and J. Maki-Paakkanen. 1995. Mutagenicity in vitro
of 3,4-dichloro-5-hydroxy-2(5H)-furanone (mucochloric acid), a chlorine disinfection by-product
in drinking water. Environ Mol Mutagen. 25:284-287. (As cited in EPA 2000e)
Ji, Y., C. Qin-Yao, W. Xiao-fei, L. Yi and L. Hong-mei. 1998. Prescreening teratogenic
potential of chlorinated drinking water disinfection by-products by using Hydra regeneration
assay. Journal of Environmental Sciences. 10(1): 110-112. (As cited in EPA 2002a)
Johnson, P.D., B.V. Dawson and S.J. Goldberg. 1998. Cardiac teratogenicity of
trichloroethylene metabolites. Journal of the American College of Cardiology. 32(2): 540-545.
(As cited in EPA, 2002a)
Kallen, B.A.J, and E. Robert. 2000. Drinking water chlorination and delivery outcome - a
registry based study in Sweden. Reprod Toxicol. 14: 303-309. (As cited in Reif et al. 2000)
Kanitz, S., Y. Franco, V. Patrone, M. Caltabellotta, E. Raffo, C. Riggi, D. Timitilli and G.
Ravera. 1996. Association between drinking water disinfection and somatic parameters at birth.
Environ. Health Perspect. 104: 516-520. (As cited in EPA 2000c)
Kato-Weinstein, J., M.K. Lingohr, G.A. Orner, B.D. Thrall and R.J. Bull. 1998. Effects of
dichloroacetate on glycogen metabolism in B6C3F1 mice. Toxicol. 130: 141-154. (As cited in
EPA 2001c)
194
-------
Katz, R., C.N. Thai, R.M. Diener, R.F. McConnell and D.E. Semonick. 1981. Dichloroacetate,
sodium: 3-month oral toxicity studies in rats and dogs. Toxicol Appl Pharmacol. 57: 273-87.
(As cited in EPA 1998c)
King, W.D., L. Dodds and A.C. Allen. 2000a. Relation between stillbirth and specific chlorination
byproducts in public water supplies. Environ Health Perspect. 108:883-886. (As cited in EPA
2002c)
King, W.D., L.D. Marrett and C.G. Woolcott. 2000b. Case-control study of colon and rectal
cancers and chlorination by-products in treated water. Cancer epidemiology, Biomarkers and
Prevention. 9:813-818. (As cited in EPA 2001 e)
Klinefelter G.R,. L.F Strader, J.D. Suarez, andN.L. Roberts. 2002. Bromochloroacetic acid
exerts qualitative effects on sperm: implications for a novel biomarker. Toxicol Sci. 68(1): 164-
73. (As cited in EPA, 2002b)
Klinefelter, G.R., Hunter, E.S., and Narotsky, M. 2001. Reproductive and Developmental
Toxicity Associated with Disinfection By-Products of Drinking Water. In: Microbial Pathogens
and Disinfection By-Products of Drinking Water, I LSI Press, 309-333. (As summarized by the
study author, personal communication.)
Klinefelter, G., L. Strader, J. Suarez, N. Roberts, M. Holmes and L. Mole. 2000.
Dibromoacetic acid, a drinking water disinfection by-product, alters male reproductive
development and fertility. Paper presented at the annual meeting for the Society for the Study of
Reproduction, Madison, Wisconsin. (As cited in EPA 2002b)
Klinefelter, G.R., Suarez, J.D., Roberts, N.L., DeAngelo, A.B. 1995. Preliminary screening for
the potential of drinking water disinfection byproducts to alter male reproduction. Reprod
Toxicol. 9(6): 571-8. (As cited in EPA 2002c)
Klotz, J.B. and L.A. Pyrch. 1999. Neural tube defects and drinking water disinfection byproducts.
Epidemiology. 10(4): 383-390. (As cited in EPA 2002a, 2002c)
Kohan, M.J., G. Huggins-Clark and S.E. George. 1998. Mutagenicity of chlorinated and
brominated acetic acids. 29th Annual Meeting of the Environmental Mutagen Society, Anaheim,
California, March 21-26, 1998. Environ Mol Mutagen. 31: 36. (As cited in EPA 2002b)
Koivusalo, M., T. Hakulinen, T. Vartianinen, E. Pukkala, J.J. Jaakkola, and J. Tuomisto. 1998.
Drinking water mutagenicity and urinary tract cancers: a population-based case-control study in
Finland. American Journal of Epidemiology. 148(7): 704-12. (As cited in EPA 2001e)
195
-------
Komulainen, H., H. Huuskonen, V-M. Kosma, S. Lotjonen, and T. Vartiainen. 1994. Toxic
effects and excretion in urine of 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) in
the rat after a single oral dose. Arch Toxicol. 68: 398-400. (As cited in EPA 2000e)
Komulainen, H., V-M. Kosma, S-L. Vaittinen, T. Vartiainen, E. Kaliste-Korhonen, S. Lotjonen,
R.K. Tuominen and J. Tuomisto. 1997. Carcinogenicity of the drinking water mutagen 3-chloro-
4-(dichloromethyl)-5-hydroxy-2(5H)-furanone in the rat. J Natl Cancer Inst. 89: 848-856. (As
cited in EPA 2000e)
Kramer, M.D., C.F. Lynch, P. Isacson and J.W. Hanson. 1992. The association of waterborne
chloroform with intrauterine growth retardation. Epidemiology. 3(5): 407-13. (As cited in EPA
1998c)
Kurokawa, Y., S. Aoki, Y. Matsushima, N. Takamura, T. Imazawa and Y. Hayashi. 1986a.
Dose-response studies on the carcinogenicity of potassium bromate in F344 rats after long-term
oral administration. J Natl Cancer Inst. 77: 977-982. (As cited in EPA 1998c, 2001b)
Kurokawa, Y., S. Takayama, Y. Konishi, Y. Hiasa, S. Asahina, M. Takahashi, A. Maekawa and
Y. Hayashi. 1986b. Long-term in vivo carcinogenicity tests of potassium bromate, sodium
hypochlorite and sodium chlorite conducted in Japan. Environ Health Perspec. 69: 221-236.
(As cited in EPA 1998c)
Kurokawa, Y.A., A. Maekawa, M. Takahashi and Y. Hayashi. 1990. Toxicity and
carcinogenicity of potassium bromate—a new renal carcinogen. Environ Health Perspect. 87:
309-35. (As cited in EPA 1998c, 2001b)
Kutom, A., N.G. Bazilinski, L. Magana and G. Dunea. 1990. Bromate intoxication:
Hairdressers'anuria. Am J Kidney Dis. 15(1): 84-85. (As cited in EPA 2001b)
Kuwahara, T., Y. Ikehara and K. Kanatsu, T. Doi, H. Nagai, H. Nakayashiki, T. Tamura and C.
Kawai. 1984. 2 Cases of potassium bromate poisoning requiring long-term hemodialysis therapy
for irreversible tubular damage. Nephron. 37:278-280. (As cited in EPA 2001b)
Laberge C., A. Lescault, and R.M. Tanguay. 1986. Hereditary tyrosinemias (type I): a new vista
on tyrosine toxicity. Adv Exp Med Biol. 206: 209-221. (As cited in EPA 2001c)
Landi, S., N.M. Handley, S.H. Warren, R.A. Pegram and D.M. DeMarini. 1999. Induction of
genetic damage in human lymphocytes and mutations in Salmonella by trihalomethanes: role of
red blood cells and GSTTI-1 polymorphism. Mutagenesis. 14: 479-482. (As cited in EPA 2002c)
196
-------
Lapinskas, P.J. and J.C. Corton. 1999. Molecular mechanisms of hepatocarcinogenic
peroxisome proliferators. In: Molecular Biology of the Toxic Response. Eds. A. Puga and K.B.
Wallace. Taylor and Francis, Philadelphia, PA. pp. 219-253. (As cited in EPA 2002a)
Lash, L.H., J.W. Fisher, J.C. Lipscomb and J.C. Parker. 2000. Metabolism of trichlorethylene.
Environ Health Perspect. 108 (Suppl 2): 177-200. (As cited in EPA 2002a)
Latendresse, J.R. and M.A. Pereira. 1997. Dissimilar characteristics of N-methyl-N-nitrosourea-
initiated foci and tumors promoted by dichloroacetic acid or trichloroacetic acid in the liver of
female B6C3F1 mice. Toxicologic Pathology. 25(5): 433-440. (As cited in EPA 2002a)
Leavitt, S.A., A.B. DeAngelo, M.H. George and J. A. Ross. 1997. Assessment of the
mutagenicity of dichloroacetic acid in lacl transgenic B6C3F1 mouse liver. Carcinogenesis.
18(11): 2101-2106. (As cited in EPA 2001c)
Le Curieux, F., F. Nesslany, T. Munter, L. Kronberg and D. Marzin. 1999. Genotoxic activity of
chlorohydroxyfuranones in the microscale micronucleus test on mouse lymphoma cells and the
unscheduled DNA synthesis assay in rat hepatocytes. Mutagenesis. 14: 457-462. (As cited in
EPA 2000e)
Lichtenberg, R., W.P. Zeller, R. Gatson and R.M. Hurley. 1989. Bromate poisoning. J. Pediatr.
114: 891-894. (As cited in EPA 2001b)
Lieber, L. 1997. Cytochrome P-4502E1: its physiological and pathological role. Physiol. Rev.
77(2):517-544. (As cited in EPA 2002c)
Linder, R.E., G.R. Klinefelter, L.F. Strader, J.D. Suarez and C.J. Dyer. 1994a. Acute
spermatogenic effects of bromoacetic acids. Fundamental and Applied Toxicology. 22: 422-430.
(As cited in EPA 2002b)
Linder, R.E., G.R. Klinefelter, L.F. Strader, J.D. Suarez, N.L. Roberts and C.J. Dyer. 1994b.
Spermatotoxicity of dibromoacetic acid in rats after 14 daily exposures. Reproductive
Toxicology. 8(3): 251-259. (As cited in EPA 2002b)
Linder, R.E., G.R. Klinefelter, L.F. Strader, M.G. Narotsky, J.D. Suarez, N.L. Roberts and S.D.
Perreault. 1995. Dibromoacetic acid affects reproductive competence and sperm quality in the
male rat. Fundam and Appl Toxicol. 28: 9-17. (As cited in EPA 2002b)
197
-------
Linder, R.E., G.R. Klinefelter, L.F. Strader, J.D. Suarez andN.L. Roberts. 1997a.
Spermatotoxicity of dichloroacetic acid. Reproductive Toxicology. 11(5): 681-688. (As cited in
EPA 2001c)
Linder, R.E., G.R. Klinefelter, L.F. Strader, D.N.R. Veeramachaneni, N.L. Roberts and J.D.
Suarez. 1997b. Histopathologic changes in the testes of rats exposed to dibromoacetic acid.
Reproductive Toxicology. 11(1): 47-56. (As cited in EPA 2002b)
Lubbers, J.R., S. Chauhan, and J.R. Bianchine. 1982. Controlled clinical evaluations of chlorine
dioxide, chlorite and chlorate in man. Environ Health Perspect. 46: 57-62. (As cited in EPA
2000c)
Lubbers, J.R., Chauhan, S., and J. R. Bianchine. 1981. Controlled clinical evaluations of chlorine
dioxide, chlorite and chlorate in man. Fundam Appl Toxicol. 1: 334-338. (As cited in EPA
2000c)
Lubbers, J.R., S. Chauhan, J.K. Miller et al. 1984. The effects of chronic administration of
chlorine dioxide, chlorite and chlorate to normal healthy adult male volunteers. J Environ Pathol
Toxicol Oncol. 5: 229-238. (As cited in EPA 2000c)
Lue, J.N., C.E. Johnson and D.L. Edwards. 1988. Bromate poisoning from ingestion of
professional hair-care neutralizer. Clinical Pharmacy. 7:66-70. (As cited in EPA 2001b)
Luft, J.C., J.B. Garges, J.R. Rockett and D.J. Dix. 2000. Male reproductive toxicity of
bromochloroacetic acid in mice. Paper presented at the annual meeting for the Society for
the Study of Reproduction, Madison, Wisconsin. Abstract. (As cited in EPA 2002b)
Mack, R.B. 1988. Round up the usual suspects. Potassium bromate poisoning. N.C. Med J.
49: 243-245. (As cited in EPA 2001b)
Mackay, J.M., V. Fox, K. Griffiths, D.A. Fox, C.A. Howard, C. Coutts, I. Wyatt and J.A. Styles.
1995. Trichloroacetic acid: investigation into the mechanism of chromosomal damage in the in
vitro human lymphocyte cytogenetic assay and the mouse bone marrow micronucleus test.
Carcinogenesis. 16(5): 1127-1133. (As cited in EPA 2002a)
Magnus, P., J.J. Jaakkola, A. Skrondal, J. Alexander, G. Becher, T. Krogh and E. Dybing. 1999.
Water chlorination and birth defects. Epidemiology. 10(5): 513-517. (Abstract only). (As cited
in Reif et al. 2000)
198
-------
Maki-Paakkanen, J., K. Jansson, and T. Vartiainen. 1994. Induction of mutation, sister
chromatid exchanges, and chromosome aberrations by 3-chloro-4-(dichloromethyl)-5-hydroxy-2-
[5H] furanone in Chinese hamster ovary cells. Mutat Res. 310:117-123. (As cited in EPA
2000e)
Maki-Paakkanen, J. and K. Jansson. 1995. Cytogenetic effects in the peripheral lymphocytes and
kidney cells of rats exposed to 3-chloro-4-(dichloromethyl)-5-hydroxy-2-[5H] furanone (MX)
orally on three consecutive days. Mutat Res. 343:151-156. (As cited in EPA 2000e)
Mather, G.G., J.H. Exon and L.D. Koller. 1990. Subchronic 90-day toxicity of dichloroacetic
acid and trichloroacetic in rats. Toxicology. 64: 71-80. (As cited in EPA 1998c)
Matsumoto, I., T. Morizono and M.M. Paparella. 1980. Hearing loss following potassium
bromate: Two case reports. Otolaryngol. Head Neck Surg. 88: 625-629. (As cited in EPA
2001b)
Mayer, P., R. Hirsch, G. Reifferscheid and K. Haberer. 1996. Development of a joint sample
pretreatment for the successive determination of organic compounds and genotoxicity in
chlorinated water. VomWasser. 86:305-320. (As cited in EPA 2002b)
McGregor, D.B., R. Martin, P. Cattanach, I. Edwards, D. McBride and W. J. Caspari. 1987.
Responses of the LS178Y tk+/tk- mouse lymphoma cell forward mutation assay: III. 72 Coded
chemicals. Environ Molec Mutagen. 12:85-154. (As cited in EPA 2002a)
Meier, J.R., Bull, R.J., Stober, J.A., Cimino, M.C. 1985. Evaluation of chemicals used for
drinking water disinfection for production of chromosomal damage and sperm-head abnormalities
in mice. Environ Mutagen. 7(2): 201-11. (As cited in EPA, 1998c)
Meier, J.R., W.F. Blasak, and R.B. Knohl. 1987. Mutagenic and clastogenic properties of 3-
chloro-4-(dichloromethyl)-5-hydroxy-2-[5H] furanone: a potent bacterial mutagen in drinking
water. Environ Mol Mutagen. 10:411-424. (As cited in EPA 2000e)
Meier, J.R., A.B. DeAngelo, F.B. Daniel, K.M. Schenck, M.F. Skelly, and S.L. Huang. 1989.
DNA adduct formation in bacterial and mammalian cells by 3-chloro-4-(dichloromethyl)-5-
hydroxy-2-[5H] furanone (MX). Environ Mol Mutagen. 14:128. (As cited in EPA 2000e)
Meier, J.R., S. Monarca, K.S. Patterson, M. Villarini, F.B. Daniel, M. Monetti and R. Pasquini.
1996. Urine mutagenicity and biochemical effects of the drinking water mutagen, 3-chloro-4-
(dichloromethyl)-5-hydroxy-2-[5H] furanone (MX), following repeated oral administration to
mice and rats. Toxicology. 110: 59-70. (As cited in EPA 2000e)
199
-------
Melnick, R.L., G.A. Boorman and V. Dellarco. 1997. Water chlorination, 3-chloro-4-
(dichloromethyl)-5-hydroxy -2(5H)-furanone (MX), and potential cancer risk. J Natl Cancer Inst.
89: 832-833. (As cited in EPA 2000e)
Melnick, R.L., M.C. Kohn, J.K. Dunnick, et al. 1998. Regenerative hyperplasia is not required for
liver tumor induction in female B6C3F1 mice exposed to trihalomethanes. Toxicol Appl
Pharmacol. 148(1): 137-147. (As cited in EPA 2002c)
Mera, N., S. Ohmori, S. Itahashi, K. Kiuchi, M. Igarashi, T. Rikihisa and M. Kitada. 1994.
Immunochemical evidence for the occurrence of Mu class glutathione S-transferase in human fetal
livers. J Biochem (Tokyo). 116: 315-320. (As cited in EPA 2002c)
Miller, R.G., F.C. Kopfler, L.W. Condie, M.A. Pereira, J.R. Meier, H.P. Ringhand, M. Robinson
and B.C. Casto. 1986. Results of toxicological testing of Jefferson Parish pilot plant samples.
Environ Health Perspect. 69: 129-139. (As cited in EPA 2000c)
Mobley, S.A., D.H. Taylor, R.D. Laurie, et al. 1990. Chlorine dioxide depresses T3 uptake and
delays development of locomotor activity in young rats. In: Jolley, R. L., et al., eds. Water
chlorination: chemistry, environmental impact and health effects, vol. 6. Chelsea, MI: Lewis
Publications, pp. 347-358. (As cited in EPA 2000c)
Moore, G.S., Calabrese, E.J., Leonard, D.A. 1980. Effects of chlorite exposure on conception
rate and litters of A/J strain mice. Bull Environ Contam Toxicol. 25: 689-696. (As cited in EPA
2000c)
Moore, G.S., Calabrese, E.J. 1982. Toxicological effects of chlorite in the mouse. Environ
Health Perspect. 46: 31-37. (As cited in EPA 2000c)
Moore, M.M. and K. Harrington-Brock. 2000. Mutagenicity of trichloroethylene and its
metabolites: Implications for the risk assessment of trichloroethylene. Environ Health Perspect.
108 (Suppl 2): 215-23. (As cited in EPA 2002a)
Mortelmans, K., S. Haworth, T. Lawlor et al. 1986. Salmonella mutagenicity tests: II. Results
from the testing of 270 chemicals. Environ Mutagen. 8: 1-119. (As cited in EPA 2002a)
Moser, V.C., P.M. Phillips, K.L. McDaniel and R.C. MacPhail. 1999. Behavioral evaluation of
the neurotoxicity produced by dichloroacetic acid in rats. Neurotoxicol and Teratol. 21(6): 719-
731. (As cited in EPA 2001 c)
200
-------
Murray, F.J., Schwetz, B.A., McBride, J.G., Staples, R.E. 1979. Toxicity of inhaled chloroform
in pregnant mice and their offspring. Toxicol Appl Pharmacol. 50:515-22. (As cited in EPA
2001a)
Nakano, K., S. Okada, S. Toyokuni and O. Midorikawa. 1989. Renal changes induced by
chronic oral administration of potassium bromate or ferric nitrilotriacetate in Wistar rats. Jpn.
Arch Intern Med. 36:41-47. (As cited in EPA 2001b)
Narotsky, M.G., B.T. Hamby, D.S. Best and E.S. Hunter. 1996. In vivo developmental effects
of dibromoacetic acid (DBA) and dichloroacetic acid (DCA) in mice (Abstract). Teratology.
53(2): 96-97. (As cited in EPA 2002b)
Narotsky, M.G., R.A. Pegram and R.J. Kavlock. 1997a. Effect of dosing vehicle on the
developmental toxicity of bromodichloromethane and carbon tetrachloride in rats. Fundam Appl
Toxicol. 40: 30-6. (As cited in EPA 2002c)
Narotsky, M.G., B.T. Hamby and D.S. Best. 1997b. Developmental effects of dibromoacetic
acid (DBA) in a Segment II study in mice. Teratology. 55(1):67.
NRC. 1994. National Research Council. Science and Judgement in Risk Assessment. National
Academy Press. (As cited in EPA, 1998c)
NRC. 1990. National Research Council. Health effects of exposure to low levels of ionizing
radiation. National Research Council. BEIR V. Washington: National Academy Press. (As cited
in EPA 1999)
NRC. 1983. National Research Council. Red Book. National Academy Press. (As cited in
EPA, 1998c)
NTP. 2000a. National Toxicology Program. Water disinfection byproducts (bromoacetic acid).
Available on-line at http://ntp-server.niehs.nih.gov/htdocs/Results_Status/Resstatb/
M920034.html (As cited in EPA 2002b)
NTP. 2000b. National Toxicology Program. Water disinfection byproducts (bromochloroacetic
acid). Available on-line at http://ntp-
server.niehs.nih.gOv/htdocs/Results_Status/Resstatw/M980085.html. (As cited in EPA 2002b)
NTP. 2000c. National Toxicology Program. Water disinfection byproducts (dibromoacetic
acid). Available on-line at http://ntp-server.niehs.nih.gov/htdocs/Results_Status/Resstatw/
M960093.html. (As cited in EPA 2002b)
201
-------
NTP. 1999. National Toxicology Program. Final range-finding report. Immunotoxicity of
dibromoacetic acid in female B6C3F1 mice. Submitted to the National Toxicology Program.
Final Report. Richmond, VA: White. K.L. and Munson, J. A. Medical College of Virginia. (As
cited in EPA, 2002b)
NTP. 1998a. National Toxicology Program. Final Report on the short reproductive and
developmental toxicity of bromodichloromethane (CAS No. 75-27-4) administered in drinking
water to Sprague-Dawley rats. Research Triangle Park, NC: U.S. Department of Health and
Human Services. Publication No. NTIS/PB99-111262. (As cited in EPA 2002c)
NTP. 1998b. National Toxicology Program. Bromochloroacetic acid: Short term reproductive
and developmental toxicity study when administered to Sprague-Dawley rats in the drinking
water. Submitted to the National Toxicology Program. Final Report. Gaithersburg, Maryland:
R O W. Sciences, Inc. R.O.W. Sciences Study No. 7244-301; NTP/NIEHS No.:
NO1-ES-75409; NTP-RDGT No.: 96-001. Also available as NTIS# PB98-172414. http://ntp-
server.niehs.nih.gov/htdocs/RDGT-studies/RDGT96001.html. (As cited in EPA 2002b)
NTP. 1996. National Toxicology Program. Final report on the short term reproductive and
developmental toxicity of chlorodibromomethane (CAS No. 124-48-1) administered in drinking
water to Sprague-Dawley rats. Research Triangle Park, NC: U.S. Department of Health and
Human Services. NTIS/PB97-111728. (As cited in EPA, 1998c)
NTP. 1992a. National Toxicology Program. NTP technical report on the toxicology and
carcinogenesis studies of chlorinated and chloraminated water in F344/N rats and B6C3F1 mice
(drinking water studies). NTP TR 392, National Institutes of Health. (As cited in EPA, 1998c)
NTP. 1992b. National Toxicology Program. NTP technical report on the toxicology and
carcinogenesis studies of monochloroacetic acid (CAS No. 79-11-8) in F344/N rats and B6C3F1
mice (gavage studies). NTP TR 396. NIH Publication No. 90-2851. (As cited in EPA 1998c,
2002a)
NTP. 1989a. National Toxicology Program. Bromoform: reproduction and fertility assessment
in Swiss CD-I mice when administered by gavage. Research Triangle Park, NC: U.S.
Department of Health and Human Services. NTP-89-068. (As cited in EPA, 1998c)
NTP. 1989b. National Toxicology Program. Toxicology and carcinogenesis studies of
tribromomethane (bromoform) in F344/N rats and B6C3F1 mice (gavage studies). Research
Triangle Park, NC: U.S. Department of Health and Human Services. Technical Report Series
No. 350. (As cited in EPA, 1998c; WHO, 2000)
202
-------
NTP. 1988. National Toxicology Program. Chloroform reproduction and fertility assessment in
CD-I mice when administered by gavage. Report by Environmental Health Research and Testing,
Inc., Lexington, KY to National Toxicology Program, NTP- 89-018. NTIS PB89-148639. (As
cited in EPA, 1998c)
NTP. 1987. National Toxicology Program. Toxicity and carcinogenesis studies of
bromodichloromethane (CAS No. 75-27-4) in F344/N rats and B6C3F1 mice (gavage studies).
Research Triangle Park, NC: U.S. Department of Health and Human Services. Technical Report
Series No. 321. (As cited in EPA, 1998c)
NTP. 1985. National Toxicology Program. Toxicology and carcinogenesis studies of
chlorodibromomethane in F344/N rats and B6C3F1 mice (gavage studies). Research Triangle
Park, NC: U.S. Department of Health and Human Services. Technical Report Series No. 282.
(As cited in EPA, 1998c)
Nelson, M.A. and R.J. Bull. 1988. Induction of strand breaks in DNA by trichloroethylene and
metabolites in rat and mouse liver in vivo. Toxicol Appl Pharmacol. 94: 45-54. (As cited in EPA
2002a)
Nishikawa, A., F. Furukawa, I. S. Lee, K. I. Kasahara, Z. Tanakamaru, H. Nakamura, M.
Miyauchi, N. Kinae and M. Hirose. 1999. Promoting effects of 3-chloro-4-(dichloromethyl)-5-
hydroxy-2(5H)-furanone on rat glandular stomach carcinogenesis initiated with N-methyl-
N'-nitro-N-nitrosoguanidine. Cancer Res. 59:2045-2049. (As cited in EPA 2000e)
Nuckols, J.R., L. Stallones and J. Reif, 1995. Evaluation of the use of a geographic information
system in drinking water epidemiology. Assessing and Managing Risks from Drinking Water
Contamination: Approaches and Applications. IAHS Publ. 233. pp. 111-22. (As cited in EPA
1998c)
Oesterheld, J.R. 1998. A review of developmental aspects of cytochrome P-450. J Child
Adolesc Psychopharmacol. 8(3): 161-174. (As cited in EPA 2000d)
Orme, J., D.H. Taylor, R.D. Laurie, et al. 1985. Effects of chlorine dioxide on thyroid function
in neonatal rats. J Toxicol Environ Health. 15: 315-322.44. (As cited in EPA 2000c)
Parker, W.A. and J.R. Barr. 1951. Potassium bromate poisoning. Br Med J. 1:1363. (As cited
in EPA 2001b)
203
-------
Parrish, J.M., E.W. Austin, D.K. Stevens, D.H. Kinder and R.J. Bull. 1996. Haloacetate-induced
oxidative damage to DNA in the liver of male B6C3F1 mice. Toxicology. 110: 103-111. (As
cited in EPA 2001c, 2002a, 2002b)
Pereira, M.A. 1996. Carcinogenic activity of dichloroacetic acid and trichloroacetic acid in the
liver of female B6C3F1 mice. Fundam Appl Toxicol. 31:192-199. (As cited in EPA 2001c,
2002a)
Pereira, M.A. and J.B. Phelps. 1996. Promotion by dichloroacetic acid and trichloroacetic acid
of N-methyl-N-nitrosourea-initiated cancer in the liver of female B6C3F1 mice. Cancer Lett.
102: 13 3 -141. (As cited in EPA 2001 c, 2002a)
Potter, C. L., L.W. Chang, A.B. DeAngelo and F.B. Daniel. 1996. Effects of four trihalomethanes
on DNA strand breaks, renal hyaline droplet formation, and serum testosterone in male F-344
rats. Cancer Lett. 106(2): 235-242. (As cited in EPA 2002c)
Quick, C.A., R.A. Chole and S.M. Mauer. 1975. Deafness and renal failure due to potassium
bromate poisoning. Arch Otolaryngol. 101: 494-495. (As cited in EPA 2001b)
Randall, J.L., S.A. Christ, P. Horton Perez, G.A. Nolen, E.J. Read and M.K. Smith. 1991.
Developmental effects of 2-bromoacetic acid in the Long Evans rat. Teratology. P122: 454.
(As cited in EPA 2002b)
Reif, J.S., A. Bachand and M. Andersen. 2000. Reproductive and developmental effects of
disinfection by-products. Final report, October 31, 2000. Prepared for Bureau of Reproductive
and Child Health, Health Canada.
Rendic, S. and F.J. Di Carlo. 1997. Human cytochrome P-450 enzymes: a status report
summarizing their reactions, substrates, inducers, and inhibitors. Drug Metab Rev. 29(1-2): 413-
580. (As cited in EPA 2000d)
Renwick, A.G., and N.R. Lazarus. 1998. Human variability and noncancer risk assessment - an
analysis of the default uncertainty factor. Regul Toxicol Pharmacol. 27(1 pt 1): 3-20.
Richard, A.M. and E.S. Hunter, III. 1996. Quantitative structure-activity relationships for the
developmental toxicity of haloacetic acids in mammalian whole embryo culture. Teratology. 53:
352-360. (As cited in EPA 2002b)
204
-------
Richmond, R.E., J.H. Carter, H.W. Carter, F.B. Daniel, and A.B. DeAngelo. 1995.
Immunohistochemical analysis of dichloroacetic acid (DCA)-induced hepatocarcinogenesis in
male Fischer (F244) rats. Cancer Lett. 25:67-76
Ruddick, J. A., D.C. Villeneuve and I. Chu. 1983. A teratological assessment of four
trihalomethanes in the rat. J Environ Sci Health. 1318(3): 333-49. (As cited in EPA, 1998c)
Saillenfait, A.M., I. Langonne and J.P. Sabate. 1995. Developmental toxicity of trichloroethylene,
tetrachloroethylene and four of their metabolites in rat whole embryo culture. Arch Toxicol. 70:
71-82. (As cited in EPA, 1998c)
Sasaki, Y. F., E. Nishidate, F. Izumiyama, M.Watanabe-Akanuma, N. Kinae, N. Matsusaka and
S. Tsuda. 1997. Detection of in vivo genotoxicity of 3-chloro-4-(dichloromethyl)-5-hydroxy-2-
[5H] furanone (MX) by the alkaline single cell electrophoresis (comet) assay in multiple mouse
organs. Mutat Res. 393:47-53. (As cited in EPA 2000e)
Sasaki, Y.F. and N. Kinae. 1995. Co-clastogenic effects of water pollutants. Environmental
Mutagen Research Communications. 17: 65-74. (As cited in EPA 2002b)
Savitz, D.A., K.W. Andrews and L.M. Pastore. 1995. Drinking water and pregnancy outcome in
central North Carolina: source, amount, and trihalomethane levels. Environ Health Perspect.
103(6): 592-6. (As cited in EPA 2002c)
Sawada, M., T. Sofuni and M. Ishidate, Jr. 1987. Cytogenetic studies on 1,1-dichloroethylene
and its two isomers in mammalian cells in vitro and in vivo. Mutat Res. 187(3): 157-163. (As
cited in EPA 2002a)
Schenkman, M.D., K.E. Tummel and L.V. Favreau. 1989. Physiological and pathological
alterations in rat hepatic cytochrome P-450. Drug Met Rev. 20(2-4): 557-84. (As cited in EPA,
1998b)
Scheuplein, R., G. Charnley, and M. Dourson. 2002. Differential sensitivity of children and
adults to chemical toxicity. I. Biological basis. Regul Toxicol Pharmacol. 35(3): 429-447.
Schwetz, B.A., B.J.K. Leong and P.J. Gehring. 1974. Embryo- and fetotoxicity of inhaled
chloroform in rats. Toxicol Appl Pharmacol. 28: 442-51. (As cited in EPA, 1998c)
Scully, F.E. and W.N. White. 1991. Reactions of chlorine, monochloramine in the G.I. tract.
Environ Sci Technol. 25(5): 820-8. (As cited in EPA, 1998c)
205
-------
Sekihashi, K., A. Yamamoto, Y. Matsumura, S. Ueno, M. Watanabe-Akanuma, F. Kassie, S.
Knasmuller, S. Tsuda, and Y.F. Sasaki. 2002. Comparative investigation of multiple organs of
mice and rats in the comet assay. Mutat Res. 517(l-2):53-75. (As cited in EPA, 2002c)
Shaw, G.M., L.H. Malcoe, A. Milea and S.H. Swan. 1991. Chlorinated water exposures and
congenital cardiac anomalies. Epidemiology. 2(6): 459-60. (As cited in Reif et al. 2000)
Shaw, G.M., S.H. Swan, J.A. Harris and L.H. Malcoe. 1990. Maternal water consumption
during pregnancy and congenital cardiac anomalies. Epidemiology. 1: 206-211. (As cited in Reif
et al. 2000)
Simmon, V.F. and R.G. Tardiff. 1978. The mutagenic activity of halogenated compounds found
in chlorinated drinking water. In: Water chlorination: Environmental Impacts and Health
Effects. Ann Arbor, MI: Ann Arbor Science, pp. 2:417-431. (As cited in EPA 2002c)
Smith, M. K., J.L. Randall, E.J. Read and J.A. Stober. 1989. Teratogenic activity of
trichloroacetic acid in the rat. Teratology. 40: 445-51. (As cited in EPA, 1998c)
Smith, M.K., J.L. Randall, E.J. Read and J.A. Stober. 1990. Developmental effects of
chloroacetic acid in the Long-Evans rat (abstract PI64). Teratology. P164: 593. (As cited in
EPA 2002a)
Smith, M.K., J.L. Randall, E.J. Read and J.A. Stober. 1992. Developmental toxicity of
dichloroacetate in the rat. Teratology. 46(3): 217-23. (As cited in EPA, 1998c)
So, B. and R.J. Bull. 1995. Dibromoacetate (DBA) acts as a promoter of abnormal crypt foci in
the colon of F344 rats. Toxicologist. 15(1): 232. (As cited in EPA 2002b)
Sofuni, T., M. Honma, M. Hayashi, H. Shimada, N. Tanaka, S. Wakuri, T. Awogi, K. I.
Yamamoto, Y. Nishi and M. Nakadate. 1996. Detection of in vitro clastagens and spindle poisons
by the mouse lymphoma assay using the micro well method: interim report of an international
collaborative study. Mutagenesis. 11(4): 349-355. (As cited in EPA 2002c)
Song, B.J., Gelboin, H.V., Park, S.S., Yang, C.S., Gonzalez, F.J. 1986. Complementary DNA
and protein sequences of ethanol-inducible rat and human cytochrome P-450s. J Biol Chem.
261(35): 16689-97. (As cited in EPA, 1998b)
Stacpoole, P.W., G.N. Henderson, Z. Yan and M.O. James. 1998. Clinical pharmacology and
toxicology of dichloroacetate. Environ Health Perspect. 106(suppl. 4): 989-994. (As cited in
EPA 2001c, 2002a)
206
-------
Stauber, A.L, A.B. DeAngelo and R.J. Bull. 1995. Different modes of action of chlorinated and
brominated haloacetates. Toxicol Sci. 15(Suppl. 1): 232. (As cited in EPA 2002b)
Stauber, A.J. and R.J. Bull. 1997. Differences in phenotype and cell replicative behavior of
hepatic tumors induced by dichloroacetate (DCA) and trichloroacetate (TCA). Toxicol Appl
Pharmacol. 144(2): 235-46. (As cited in EPA 2002a, 2002b)
Stratton, C.E., W.E. Ross and S. Chapman. 1981. Cytotoxity and deoxyribonucleic acid damage
associated with bromoacetate. Biochem Pharmacol. 30: 1497-1500. (As cited in EPA 2002b)
Suh, D.H., M.S. Abdel-Rahman and R.J. Bull. 1983. Effect of chlorine dioxide and its
metabolites in drinking water on fetal development in rats. J Appl Toxicol. 3: 75-79. (As cited in
EPA 2000c)
Swan, S.H., K. Waller, B. Hopkins, G. Windham, L. Fenster, C. Schaefer and R.R. Neutra. 1998.
A prospective study of spontaneous abortion: relation to amount and source of drinking water
consumed in early pregnancy. Epidemiology. 9(2): 126-33. (As cited in EPA, 1998c)
Tanaka, E. 1998. In vivo age-related changes in hepatic drug-oxidizing capacity in humans. J.
Clinical Pharm. Therapeutics. 23(4): 247-255. (As cited in EPA 2000d)
Tanguay, R.M., R. Jorquera, J. Poudrier, et al. 1996. Tyrosine and its catabolites: From disease
to cancer. Acta Biochim Pol. 43: 209-216. (As cited in EPA 2001c)
Tao, L., K. Li, P.M. Kramer and M.A. Pereira. 1996. Loss of heterozygosity on chromosome 6
in dichloroacetic acid and trichloroacetic acid-induced liver tumors in female B6C3F1 mice.
Cancer Lett. 108: 257-261. (As cited in EPA 2002a)
Tao, L., P.M. Kramer, R. Ge, and M.A. Pereira. 1998. Effect of dichloroacetic acid and
trichloroacetic acid on DNA methylation in liver and tumors of female B6C3F1 mice.
Toxicological Sciences. 43:139-144. (As cited in EPA 2002a)
Taylor, D.H. and R.J. Pfohl. 1985. Effects of chlorine dioxide on the neurobehavioral
development of rats. In: Jolley, R.L., et al., eds. Water chlorination: chemistry, environmental
impact and health effects, vol. 6. Chelsea, MI: Lewis Publications, pp. 355-364. (As cited in EPA
2000c)
Teramoto, S., K. Takahashi, M. Kikuta and H. Kobayashi. 1998. Potential teratogenicity of 3-
chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) in micromass in vitro test. J Toxicol
Environ Health. 53: 607-614. (As cited in EPA, 2000e)
207
-------
Theiss, J.C., G.D. Stoner, M.B. Shimkin, and E.K. Weisburger. 1977. Test for carcinogenicity of
organic contaminants of United States drinking waters by pulmonary tumor response in strain A
mice. Cancer Res. 37(8): 2717-2720. (As cited in EPA 2002c)
Thompson, D.J., S.D. Warner and V.B. Robinson. 1974. Teratology studies on orally
administered chloroform in the rat and rabbit. Toxicol Appl Pharmacol. 29: 348-57. (As cited in
EPA 1998c, 2001a)
Thornton-Manning, J.R., J.C. Seely and R.A. Pegram. 1994. Toxicity of bromodichloromethane
in female rats and mice after repeated oral dosing. Toxicology. 94(1-3): 3-18. (As cited in EPA,
2002c)
Tobe, M., Y. Suzuki, K. Aida, H. Yoshimoto, K. Takada, J. Kadoma, K. Kobayashi, Y. Vehida,
K. Anda, Y. Nakaji, Y. Suzuki and M. Sato. 1982. Studies on the chronic oral toxicity of
tribromomethane, dibromochloromethane, and bromodichloromethane. Unpublished intraagency
report to the National Institute of Hygienic Sciences. Tokyo, Japan: Tokyo Medical and Dental
University. (As cited in EPA 2002c)
Tomasi, A., E. Albano, F. Biasi, T.F. Slater, V. Vannini and M.V. Diazano. 1985. Activation of
chloroform and related trihalomethanes to free radical intermediates in isolated hepatocytes and in
the rat in vivo as detected by the ESR-Spin-trapping technique. Chem. Biol. Interact. 55:303-
316. (As cited in EPA 2002c)
Tong, Z., P.G. Board and M.W. Anders. 1998a. Glutathione transferase Zeta catalyzes the
oxygenation of the carcinogen DCA to glyoxylic acid. Biochem J. 331(2): 371-374. (As cited
in EPA 2002a)
Tong, Z., P.G. Board and M.W. Anders. 1998b. Glutathione transferase zeta-catalyzed
biotransformation of dichloroacetic acid and other alpha-haloacids. Chem Res Toxicol. 11:
1332-1338. (As cited in EPA 2002a)
Toth, G.P., R.E. Long, T.S. Mills, et al. 1990. Effects of chlorine dioxide on the developing rat
brain. J Toxicol Environ Health. 31: 29-44. (As cited in EPA 2000c)
Toth, G.P., K.C. Kelty, E.L. George, E.J. Read, and M.K. Smith. 1992. Adverse male
reproductive effects following subchronic exposure of rats to sodium dichloroacetate. Fundam
Appl Toxicol. 19(1): 57-63. (As cited in EPA 2001c)
208
-------
Tuthill, R.W., R.A. Giusti, G.S. Moore and E.J. Calabrese. 1982. Health effects among newborns
after prenatal exposure to C102 -disinfected drinking water. Environ Health Perspect. 46: 39-45.
(As cited in EPA 2000c)
Ueno, T. and F.J. Gonzalez. 1990. Transcriptional control of the rat hepatic CYP2E1 gene.
Molecular and Cellular Biology. 10(9): 4495-4505. (As cited in EPA 1998b)
Umemura, T., K. Sai, A. Takagi, R. Hasegawa, and Y. Kurokawa. 1993. A possible role for cell
proliferation in potassium bromate (KBr03) carcinogenesis. J Cancer Res Clin Oncol. 119:
463-469. (As cited in EPA 2001b)
Umeno, M., B.J. Song, C. Kozak, H.V. Gelboin and F.J. Gonzalez. 1988. The rat P450IIE1
gene: Complete intron and exon sequence, chromosome mapping, and correlation of
developmental expression with specific 5' cytosine demethylation. J Biol Chem. 263(10): 4956-
62. (As cited in EPA 1998b)
Vaittinen, S-L., H. Komulainen, V-M. Kosma, A. Julkunen, J. Maki-Paakkanen, K. Jansson, T.
Vartiainen and J. Tuomisto. 1995. Subchronic toxicity of 3-chloro-4-(dichloromethyl)-5-
hydroxy-2(5H)-furanone (MX) in Wistar rats. Food Chem Toxicol. 33:1027-1037. (As cited in
EPA 2000e)
Varma, M.M., F.R. Ampy, K. Verma and W.W. Talbot. 1988. In vitro mutagenicity of water
contaminants in complex mixtures. J Appl Toxicol. 8(4): 243-248. (As cited in EPA 2002c)
Veeramachaneni, D.N.R., T.T. Higuchi, J.S. Palmer and C.M. Kane. 2000. Dibromoacetic acid,
a disinfection by-product in drinking water, impairs sexual function and fertility in male rabbits.
Paper presented at the annual meeting for the Society for the Study of Reproduction, Madison,
Wisconsin. (As cited in EPA 2002b)
Vetter, C.M., J.E. Miller, L.M. Crawford, M.J. Armstrong, J.H. Clair, M.W. Conner, L.D. Wise
and T.R. Skopek. 1998. Comparison of motility and membrane integrity to assess rat sperm
viability. Reproductive Toxicology. 12(2): 105-114. (As cited in EPA, 2002b)
Vieira, I., M. Sonnier and T. Cresteil. 1996. Developmental expression of CYP2E1 in human
liver. Hypermethylation control of gene expression during the neonatal period. Eur J Biochem.
238: 476-83. (As cited in EPA, 1998b)
Walgren, J.E., D.T. Kurtz, and J.M. McMillan. 2000. The effect of the trichloroethylene
metabolites trichloroacetate and dichloroacetate on peroxisome proliferation and DNA synthesis
in cultured human hepatocytes. Cell Biol Toxicol. 16(4): 257-73. (As cited in EPA, 2002a)
209
-------
Waller, K., S.H. Swan, G.C. Windham, and L. Fenster. 2001. Influence of exposure assessment
methods on risk estimates in an epidemiologic study of total trihalomethane exposure and
spontaneous abortion. J Expo Anal Environ Epidemiol. 11 (6):522-31. (As cited in EPA 2002c)
Waller, K., S.H. Swan, G. Delorenze and B. Hopkins. 1998. Trihalomethanes in drinking water
and spontaneous abortion. Epidemiology. 9(2): 134-40.
Ward, K.W., E.H. Rogers and E.S. Hunter, III. 2000. Comparative pathogenesis of haloacetic
acid and protein kinase inhibitor embryotoxicity in mouse whole embryo culture. Toxicological
Sciences. 53: 118-126. (As cited in EPA 2001c, 2002a, 2002b)
Ware, G.W., ed. 1989. Review of Environmental Contaminants and Toxicology. U.S.
Environmental Protection Agency, Office of Drinking Water Health Advisories. Springer-Verlag,
184 pp. (As cited in EPA, 1998c)
Warshaw, B.L., M.C. Carter, L.C. Hymes, B.S. Bruner and A.P. Rauber. 1985. Bromate
poisoning from hair permanent preparations. Pediatrics. 76(6): 975-978. (As cited in EPA
2001b)
Watanabe, T., T. Abe, M. Satoh, Y. Oda, T. Takada and T. Yanagihara. 1992. Two children
with bromate intoxication due to ingestion of the second preparation for permanent hair waving.
Acta Paediatr Jpn. 34(6): 601-605. (As cited in EPA 2001b)
WHO. 2000. World Health Organization, International Programme on Chemical Safety (IPCS).
Environmental Health Criteria 216: Disinfectants and Disinfectant By-products.
Windham, G.C., K. Waller, M. Anderson, L. Fenster, P. Mendola, and S. Swan. 2003.
Chlorination by-products in drinking water and menstrual cycle function. Env. Health Perspect.
Available at http://dx.doi.org/.
Wolf, G.W. and L. Kaiser. 1996. Final report sodium bromate: short term reproductive and
development toxicity study when administered to sprague dawley rats in the drinking water.
Submitted to National Toxicology Program, Research Triangle Park, North Carolina.
NTP/NIEHS NO. NOI-ES. (As cited in EPA 2001b)
Yang, C-Y., B-H. Cheng, S-S. Tsai, T-N. Wu, M-C. Lin, and K-C. Lin. 2000. Association
between chlorination of drinking water and adverse pregnancy outcome in Taiwan. Environ
Health Perspect. 108: 765-768. (As cited in Reif et al. 2000)
210
-------
Yang, C.Y., H.F. Chiu, M.F. Cheng, and S. S. Tsai. 1998. Chlorination of drinking water and
cancer mortality in Taiwan. Environ Res. 78(1): 1-6. (As cited in EPA 2001 e)
Zeiger, E. 1990. Mutagenicity of 42 chemicals in Salmonella. Environ Mol Mutagen Suppl.
16(Supplement 18): 32-54. (As cited in EPA 2002c)
Zenick, H., E.D. Clegg, S.D. Perreault, G.R. Klinefelter and L.E. Gray. 1994. Assessment of
male reproductive toxicity: a risk assessment approach. In: A.W. Hayes, ed. Principles and
methods of toxicology, 3rd edition. New York: Raven Press, Ltd. pp 937-988. (As cited in
EPA 2002b)
211
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