United States Office of Science and Technology
Environmental Protection Health and Ecological Criteria Division
Agency Washington, DC 20460
vvEPA
of Water
HEALTH RISKS TO FETUSES, INFANTS, AND CHILDREN
(FINAL STAGE 1 D/DBP RULE)
October 29,1998
CONTRIBUTORS
Ambika Bathija, Ph.D.
Nancy Chiu, Ph.D.
Octavia Conerly
Vicki Dellarco, Ph.D.
Kris Khanna, Ph.D.
Mary Manibusan
Yogi Patel, Ph.D.
Rita Schoeny, Ph.D.
Submitted by:
Ambika Bathija, Ph.D.
Health and Ecological Criteria Division
Office of Science and Technology
Office of Water
U.S. Environmental Protection Agency
401 M Street, SW
Washington, D.C. 20460
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CONTENTS
Abbreviations iii
Plain Language Summary of Evaluation of Children's Risk iv
1. Introduction _. I
1.1. Risk to Children ". 1
1.2. Risk Assessment Methods 2
1.3. Maximum Contaminant Level Goal and
Maximum Residual Disinfectant Level Goal 4
1.4. Determining Risk to Children .. '. 6
1.5. Summary and Conclusions 7
2. Estimates of Risk to Children From Stage 1
Disinfectants/Disinfectant Byproducts .9
2.1. Chlorinated Drinking Waters 9
2.2. Trihalomethanes 10
2.2.1. Chloroform 10
.2.2.2. Brominated Trihalomethanes .' 17
2.2.3. Haloacetic Acids 23
2.2.4. Bromate 28
2.2.5. Chlorite/Chlorine Dioxide 29
2.2.6. Chlorine 31
2.2.7. Chloramines : .. 33
3. Summary and Conclusions : 35
4. References '.-..'. 36
Table 1. Disinfectant Byproducts and Their MCLGs Cited in the Final Stage 1 D/DBP Rule .. 7
Table 2. Disinfectants and Their MRDLGs Cited in the Final Stage 1 D/DBP Rule 7
Table 3. Comparison of Toxicity Endpoints ; .8
EPA OW/OST/HECD ii Health Risks to Fetuses, Infants, and Children
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ABBREVIATIONS
BDCM bromodichloromethane
BMD benchmark dose
DBCM dibromochloromethane
DBF disinfectant byproduct
DCA dichloroacetic acid
D/DBP disinfectant and disinfectant byproduct
DWEL drinking water equivalent level
LOAEL lowest-observed-adverse-effect level
MCLG maximum contaminant level goal
MCL maximum contaminant level
MF modifying factor
MRDLG maximum residual disinfectant level goal
NOAEL no-observed-adverse-effect level
OR odds ratio
RfD reference dose
TCA trichloroacetic acid
THM , trihalomethane
TTHM total THM
UF uncertainty factor
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PLAIN LANGUAGE SUMMARY OF EVALUATION OF CHILDREN'S RISK
Executive Order 13045: Protection of Children from Environmental Health Risks and
Safety Risks
1. Does E.0.13045 apply to this rule?
This final rule is not subject to E.G. 13045, entitled "Protection of Children from Environmental
Health Risks and Safety Risks" (62 FR 19885, April 23, 1997), because the environmental health
or safety risks addressed by this action do not have a disproportionate effect on children.
2. Children's Health Protection
In accordance with section 5(501), the Agency has evaluated the environmental health or safety
effects of disinfectants and disinfectant byproducts (DBFs) on children. We conclude that the
Maximum Contaminant Level Goals (MCLGs) are protective of children. For some chemicals
this is because the MCLG was based on a study of reproductive effects or effects on developing
organisms. For the other chemicals it was determined that the toxic effects were not more likely
to occur in children or to affect children at lower doses than would affect adults. In making'these
determinations we reviewed all available data and asked the following questions for each
disinfectant or DBF:
/
1. Is there information which shows that the disinfectant or DBF causes effects in
the developing fetus or impairs ability to conceive and, bear children? If it causes
this type of problem will it occur at a lower dose than that which will cause other
types of effects?
2. If the disinfectant or DBF causes cancer, are children more likely to be affected by
it than are adults?
3. If the disinfectant or DBF causes some noncancer toxic effect, are children more
likely to be affected by it than are adults?
The answers to these questions can be found in this document.
The disinfection of public drinking water supplies to prevent waterbome disease is the
most successful public health program in U.S. history. However, numerous chemical
byproducts (DBFs) result from the reaction of chlorine and other disinfectants with naturally
occurring organic and inorganic material in source water, and these may have potential health
risks. Thus, maximizing health protection for sensitive subpopulations requires balancing risks
to achieve the recognized benefits of controlling for waterbome pathogens while minimizing risk
of potential DBF toxicity. Human experience shows that waterbome disease from pathogens in
drinking water is a major concern for children and other subgroups (elderly, immune
compromised, pregnant women) because of their greater vulnerabilities (Gerba et al., 1996).
Based on animal studies, there is also a concern for potential risks posed by DBFs to children
and pregnant women (EPA, 1994a, 1998a).
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1. INTRODUCTION
1.1. RISK TO CHILDREN
«
In 1995, EPA's Administrator established an Agencywide 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
should consider "the effects of the contaminant on the general population and on groups within
the general population such as infants, children, pregnant women, the elderly, individuals with a
history of serious illness or other subpopulations that are identified as likely to be 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 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 is following 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, 1991).
The Office of Water is charged with ensuring that the U.S. population has clean water and
safe drinking water. This mandate covers chemical, physical, and biological water pollutants.
The disinfection of public drinking water supplies to prevent waterbome disease is the most
successful public health program in U.S. history. The diarrheal diseases most commonly
associated with waterborne infectious agents have a more severe outcome for children and
debilitated adults. Thus, the savings in lives from disinfected water has largely been in those
populations. However, chemical disinfectant byproducts (DBFs) result from the reaction of
chlorine and other disinfectants with naturally occurring organic materials in source water, and
these have potential health risks. Thus, to provide for maximum human health protection it is
necessary to balance risks from water pathogens and DBFs.
A growing body of scientific knowledge demonstrates that children may suffer
disproportionately from some environmental health risks. These risks may arise because
children's neurological, immunological, and digestive systems are still developing. In addition
children may incur greater exposure, because they eat more food, drink more fluids, and breathe
more air in proportion to their body weights than do adults. Waterbome disease is a major
concern for infants, children, and other sensitive subgroups (elderly, immune-compromised,
pregnant women) because of their greater vulnerabilities (Gerba et al., 1996). Based on animal
EPA OW/OST/HECD : I Health Risks to Fetuses, Infants, and Children
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studies, there is a concern for risks posed by DBFs to children and the developing fetus (EPA.
•1994a. 1998a).
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). The 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—Does the chemical produce adverse health effects?
• Dose-Response Assessment—How does the frequency of adverse effects change
with dose?
• Exposure Assessment—How much chemical are humans exposed to in various
environments?
• Risk Characterization—Summarizes and integrates the scientific findings of the
hazard, dose-response, and exposure assessments to determine the potential
human risk.
To evaluate the toxicity of disinfectants and disinfectant 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
i . 2. . Systemic toxicity
3. Carcinogenicity
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
elderly as a late-age-onset disorder). There are a number of developmental abnormalities of
concern, including these: spontaneous abortions, stillbirths, malformations, premature mortality,
reduced birth weight, mental retardation, and sensory loss, as well as other adverse functional
EPA OW/OST/HECD 2 Health Risks to Fetuses, Infants, and Children
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and physical effects. Developmental abnormalities are extremely common and present an
enormous burden for society.
Risk assessment methodologies have been developed to estimate the magnitude of
potential harm to humans from carcinogenic and noncarcinogenic effects. Most chemicals that
do not produce carcinogenic effects are believed to have a low dose below which no adverse,
noncarcinogenic effects occur with incidence above background. On the other hand,
carcinogenic chemicals may cause some effect at low doses or not, depending on the way in
which they affect the carcinogenic process.
For contaminants with carcinogenic potential, chemical levels are measured against a
cancer potency slope factor or a unit risk value together with an assumption of lifetime exposure
from ingestion of water. The cancer unit risk has often been derived from a linearized multistage
model with a 95% upper confidence limit providing a low-dose estimate. The interpretation of
this number when derived from animal studies is that the true risk to humans is not likely to
exceed the upper-limit estimate at,low doses and, in fact, may be lower. Excess cancer risk .
estimates have also been calculated using other models such as the one-hit, Weibull, logit, and
probit.
In 1996, EPA proposed revisions to the 1986 EPA Guidelines for Carcinogenic
Assessment (EPA, 1996). Rather than relying exclusively on tumor findings, the new Guidelines
include an expanded weight-of-evidence approach that emphasizes understanding mode of
action, conditions of expression of carcinogenicity (e.g., route and magnitude of exposure), and
consideration of all other relevant data. Dose-response assessment is now done in two steps:
appropriate models are fitted to data in the empirical range of observation in animal studies to
determine a point of departure for the second step, which is extrapolation below the observable
range. The 1996 revisions to the Guidelines include several default procedures (linear, nonlinear,
or both), rather than relying on the linear multistage (LMS) model as the only default for
extrapolation of dose-response relationships. The EPA is currently applying the principles of the
new Guidelines, consistent with the 1986 Guidelines. The proposed 1996 Guidelines are not
inconsistent with the 1986 Guidelines, and findings under the 1986 Guidelines are considered to
be based on sound science for decisionmaking.
In the quantification of noncarcinogenic effects, 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
i
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 an uncertainty factor (UF) or factor
EPA OW/OST/HECD 3 Health Risks to Fetuses, Infants, and Children
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times a modifying factor (MF). The RfD ^s calculated as follows:
_ _ ' (NOAEL or LOAEL ) ., . ,
RfD = =— = mg/kg/day.
[Uncertainty Factor(s) * Modifying Factor ]
When the data support it, a benchmark dose (BMD) may be calculated by applying an
appropriate mathematical curve-fitting procedure. The BMD can then be used as a NOAEL
estimate. The NOAEL, LOAEL, or BMD is divided by a UF times an MF to calculate the RfD..
. Selection of the UF to be employed in the calculation of the RfD/RfC is based on professional
judgment, which considers the entire database of toxicologic effects for the chemical. To ensure
that UFs are selected and applied in a consistent manner, the EPA (1994a) employs a
modification to the guidelines proposed by the NAS, as shown in the box on the next page (NAS
1977, 1980).
1.3. MAXIMUM CONTAMINANT LEVEL GOAL AND MAXIMUM RESIDUAL
DISINFECTANT LEVEL GOAL
Maximum contaminant level goals (MCLGs) are nonenfprceable health goals. They are
set at concentration levels at which no known or anticipated adverse effects on the health of
persons occur, and which includes an adequate margin of safety. Establishment of an MCLG for
each 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
(see the proposed rule at 59 FR 38677 for a detailed discussion of the process for establishing
MCLGs). They can be based on an RfD for general systemic effects, on a quantitative estimate
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
proposed rule for disinfectants 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 for waterborne microbial contaminants.
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Uncertainty Factors Used in RfD/RfC Calculations
Standard Uncertainty Factors (UFs)
Use a 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.
Use an additional 10-fold factor 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. A 3-fold uncertainty factor is used for
extrapolating from inhalation studies on experimental animals to humans for the
derivation of an inhalation RfC. This difference is because dosimetric adjustments
reduce the uncertainty associated with extrapolation between experimental animals and
humans.
Use an additional 10-fold factor 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 10-fold factor 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.
Modifying Factor (MF)
\
Use professional judgment to determine another uncertainty factor (MF) that is greater
than zero and less than or equal to 10. The magnitude of the MF depends upon the
professional assessment of scientific uncertainties of the study and data base not
explicitly treated above, e.g., the completeness of the overall data base and the number
of species tested. The default value for the MF is 1.
EPA OW/OST/HECD 5 Health Risks to Fetuses, Infants, and Children
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The MCLG or MRDLG for drinking water is calculated from the RfD for a "0 kg adult
consuming 2 L of water per day and also taking into consideration the relative contribution from
drinking water. The Agency views the use of 2 L per day adult drinking water consumption to
derive the MCLG from an RfD appropriate, because it represents the 84th percentile of adult
drinking water consumption. Recent analyses show that the 90th percentile for tap water
consumption for persons in the United States aged 11-19 is 1.38 L/day. A conservative estimate
is that children may be exposed about 3.5-fold more than adults relative to their water
intake-body weight ratio (Draft Water Quality Criteria Methodology Revisions; Human Health,
August 1998). The Agency believes that the use of 2 L to calculate the MCLG provides
sufficient protection to fetuses and children.
The final Stage 1 D/DBP Rule contains MCLGs for the following disinfectant
byproducts: four trihalomethanes (THMs) (chloroform, bromodichloromethane,
dibromochloromethane, and bromoform), two haloacetic acids (dichloroacetic acid and
trichloroacetic acid), bromate, and chlorite, as shown in Table 1. The final Rule contains
MRDLGs for the following disinfectants: chlorine, chloramines, and chlorine dioxide, as
depicted in Table 2.
1.4. DETERMINING RISK TO CHILDREN
In developing this regulation, risks to sensitive subpopulations including fetuse.s 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. Do D/DBPs cause reproductive and developmental effects at doses below those
causing systemic toxicity or cancer?
2. Are fetuses and children more susceptible than adults to the systemic toxicity of
D/DBPs?
3. Are fetuses and children more susceptible than adults to cancer from D/DBPs?
For questions 2 and 3 above, mechanistic data were considered when available, that is,
for the THMs.
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Table 1. Disinfectant Byproducts and their MCLGs
Cited in the Final Stage 1 D/DBP Rule
Disinfectant Byproducts
Chloroform
Bromodichloromethane (BDCM)
Dibromochloromethane (DBCM)
Bromoform
Dichloroacetic acid (DCA)
Trichloroacetic acid (TCA)
Chlorite
Bromate
MCLG (mg/L)
0
0
0.06
0
0
0.3
0.8
0
Table 2. Disinfectants and their MRDLGs
Cited in the Final Stage 1 D/DBP Rule
Disinfectants
Chlorine
Chloramine
Chlorine dioxide
MRDLG (mg/L)
4 (as Clj)
4 (as Cy
0.8 (as ClOj)
1.5. SUMMARY AND CONCLUSIONS
This document evaluates the available data for each of the D/DBPs used for deriving the
MCLG or MRDLG, to determine if the derived MCLGs and/or MRDLGs are protective for the
fetuses and children. Table 3 summarizes the comparison of toxicity endpoints for the various
D/DBPs. As can be seen in the table, chloroform, BDCM, bromoform, DCA, and bromate are
considered likely to be carcinogenic for humans. MCLGs of zero were selected after
•
consideration of the potential carcinogenicity of these chemicals. This MCLG of zero would
protect both children and adults. The MCLG/MRDLGs for DBCM and chloramine were based
on systemic toxicity. The NOAEL/LOAEL used to derive the numbers are lower than the
NOEL/LOAELs for developmental effects; therefore, the MCLG/MRDLG would be protective
of infants and children. In the case of chlorine, the MCLG is based on systemic toxicity, because
the NOAEL of 5 mg/kg/day based on developmental effects is the highest dose tested and
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Table 3. Comparison of Toxicity Eodpoints
Disinfectant
Chloroform
Bromodichloro-
methane (BDCM)
Dibromochloro-
methane (DBCM)
Bromoform
Dichloroacetic
acid (DCA)
Trichloroacetic
acid (TCA)
Bromate
Chlorite
Chlorine dioxide
Chlorine
Chloramine
Systemic Toxicity
NOAEL'
mg/kg/day
—
—
21.00
25,00
— '
36.5
—
—
—
14.00
9.5
LOAEL1
mg/kg/day
12.90
25.00
— i
—
—
—
I
—
-r
Developmental Toxicity
NOAEL
mg/kg/day
35-50
50.00
200.00
50.00
14.00
—
15.00
3.00
3.00
5.00e
10.00
LOAEL
mg/kg/day
—
• —
—
^-
—
330.00
—
0
—
—
—
Carcino-
geniciry
probable
probable
possible
probable
probable
—
probable
—
— •
—
—
MCLG"
mg/L
0
0(Ca)
0.06 (sys tox & Ca)
0(Ca)
0 (Ca)
0.3 (devel tox)
0 (Ca)
0.8 (devel tox)
0.8 (devel tox)
4.00 (sys tox)
3.00 {sys tox)
'NOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse effect level.
"MCLG = 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.
cHighest dose tested. •.
therefore not a true NOAEL. For chlorine dioxide, chlorite, and TCA, the MCLG/MRDLGs are
calculated on data from developmental studies; hence the numbers derived would be protective
for both children and adults. It can be concluded that the MCLG/MRDLGs of all the D/DBPs in
the Stage 1 D/DBP Rule are protective of fetuses, infants, and children.
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2. ESTIMATES OF RISK TO CHILDREN FOR STAGE 1
DISINFECT ANTS/DISINFECT ANT BYPRODUCTS
Descriptions of the data available for evaluating risks to children and conclusions drawn
for each MCLG or MRDLG contained in this regulation are given in the following sections.
2.1. CHLORINATED DRINKING WATERS
This section considers studies on exposure to chlorinated drinking water rather than to
individual D/DBPs. Several epidemiology studies have reported an association between
exposure to THMs and developmental/reproductive effects (Bove et al., 1995; Nuckols et al.,
1995; Savitz et al., 1995; Waller et al., 1998). Two studies reported a relationship specifically
between developmental/reproductive toxicity and BDCM.. An epidemiological study (Kramer et
al., 1992) reported an increased risk of intrauterine growth retardation with exposure to drinking
water with BDCM concentrations greater than or. equal to 10 ug/L compared with drinking water
with undetectable BDCM concentrations (odds ration [OR] = 1.7, 95% confidence interval [CI] =
0.9-2.9). Three new reproductive epidemiology studies were published recently.
*
Klotz and Pyrch (1998) examined the potential association between neural tube defects
and certain drinking water contaminants, including some DBPs. In this case-control study, births
with neural tube defects reported to New Jersey's 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 authors reported elevated ORs, generally between 1.5 and 2.1, for the
association of neural tube defects with total THMs (TTHMsj. 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 neural tube defects in
which there were no other malformations (OR = 2.1, 95% CI = 1.1-4.0).
Two other studies investigated early-term miscarriage risk factors. The first of these
studies examined the potential association between early-term miscarriage and exposure to
THMs (Waller et al., 1998). The second study 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. They were recruited from the
Santa Clara area, the Fontana area in southern California, or the Walnut Creek area. The women
were all members of the Kaiser Permanente Medical Care Program and were offered a chance to
participate in the study when they called to arrange their first prenatal visit. In the Waller et al.
study, additional water quality information from the women's drinking water utilities were
EPA OW/OST/HECD ~9 Health Risks to Fetuses, Infants, and Children
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obtained so chat THM levels could be determined. The Swan et al. study provided no
quantitative measurements of THMs (or DBFs), and thus provided no additional information on
the risk from chlorination byproducts.
In the Waller et al. (1998) study, utilities that served the women were identified. Utilities
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. This
"first trimester TTHM level" was combined with self-reported tap water consumption to estimate
a TTHM exposure level. Exposure levels of the individual THMs (e.g., chloroform, bromoform)
were estimated in the same manner. Actual THM levels in the home tap water were not
measured.
Women with high TTHM exposure in home tap water (drinking five or more glasses per
day of cold home tap water containing at least 75 ug/L of TTHM) had an early-term miscarriage
rate of 15.7%, compared with a rate 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 ug/L of TTHM). An adjusted OR for early-term miscarriage of 1.8 (95%
confidence interval = 1.1-3.0) was determined.
Only high BDCM exposure was associated with early-term miscarriage. This was
defined as drinking five or more glasses per day of cold home tap water containing >18 ug/L
BDCM. An adjusted OR for early-term miscarriage of 3.0 (95% CI = 1.4-6.6) was determined.
2.2. TRIHALOMETHANES
2.2.1. Chloroform
Chloroform is one of the best studied DBFs and has a very extensive toxicological
database. Chloroform and its metabolites have been shown to cause liver and kidney toxicity and
tumors as its primary adverse effects.
2.2.1.1. Developmental/Reproductive Effects
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 investigated.
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Thompson et al. f 1974) studied the effects of chloroform on embryonic and fetal
development of Sprague-Dawley rats. In a range-finding study, pregnant Sprague-Dawley rats
were administered gavage doses of 79, 126, 300, 316, or 516 mg/kg/day chloroform in corn oil
on days 6-15 of gestation. The dams developed alopecia, rough hair, and eczema. Food
consumption and body weight gain were significantly decreased at 126 mg/kg/day and higher,
and 316 mg/kg/day resulted in severe maternal, toxicity and death, as well as fetotoxicity. In the
main study, groups of 25 pregnant rats (181-224 g) were gavaged with chloroform in com 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 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. The incidence of fetuses with bilateral extra lumbar ribs was significantly
(p<0.05) increased at the high dose, but 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 com 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 because three control animals aborted. Histopathology examinations
revealed no evidence of maternal 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
EPA OW/OST/HECD ' '. 11 Health Risks to Fetuses, Infants, and Children
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the mid dose <\as 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. This study is limited by the high incidence of
abortions and mortality in the control group. Due to the absence of a clear dose response, no
definitive NOAEL or LOAJEL for developmental effects can be identified for this study, although
the NOAEL appears to be in the range of 35-50 mg/kg/day.
Ruddick et al. (1983) investigated the potential 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 stemebra 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 was a clear increase in the incidence of these variations
indicating a potential developmental effect.
In a study via the inhalation route, 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 1420 mg/m3) for 7 hours/day from
gestation days 6 through 15. The numbers of pregnant rats exposed in each group were 68, 22,
23, and 3, respectively. Although the study authors attributed the low percentage of pregnancy at
the high concentration (15%) to chloroform exposure, this result can not be treatment-related, in
light of the timing of exposure. The small number of pregnant animals at 300 ppm reduced the
study sensitivity. Dams were sacrificed on gestation day 21, and fetuses delivered by caesarian
section. Numbers of live, dead, and resorbed fetuses were determined. Fetuses were weighed,
measured, and sexed. Half were examined for skeletal anomalies, and the other half were
examined for organ anomalies. High-exposure dams lost weight during exposure and ate
minimal amounts of food (approximately 1 g/rat/day); concentration-related decreases in body
weight gain and food consumption were observed at lower exposure levels. Relative liver
weights were significantly increased in dams exposed to 100 and 300 ppm at study termination,
with a significant decrease in absolute liver weight at 300 ppm. 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 (3.7 g/rat/day) than
EPA OW/OST/HECD 12 Health Risks to Fetuses, Infants, and Children
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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 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 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 Utters 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 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 doses causing maternal toxicity, and
occurred at 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. Therefore, the RfD and MCLG based on a LOAEL of 12.9
mg/kg/day for liver effects would be sufficiently protective.
EPA OW/OST/HECD \ 13 Health Risks to Fetuses, Infants, and Children
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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., Fl animals). In the first phase, mice were
administered chloroform by gavage in com 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 Fl 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 Fl
females. No NOAEL can be identified for this effect, because the low- and mid-dose groups .
were not evaluated histopathologically. However, no adverse effects on fertility or reproduction
were found.
2.2.1.2. Systemic Effects
Numerous aninial 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 (and 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), and will not be discussed in this paper. 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, 1998a). To explore the issue of whether fetuses or children are at increased cancer
risk compared with adults, the mode of carcinogenic action of chloroform was examined.
Several issues (e.g., differences in rate of cell proliferation and metabolism in children versus
adults) were explored concerning susceptibility of fetuses or children in the context of what is
known about the carcinogenic mode of action of chloroform. A substantial body of data
EPA OW/OST/HECD : 14 Health Risks to Fetuses, Infants, and Children
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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).
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, phosgene and HC1, which produce
cytotoxicity-(cell death) and regenerative hyperplasia (EPA, 1998b; ILSI, 1997). This process
may lead to tumor development if sustained. Given that oxidative metabolism is key to the
carcinogenic potential of chloroform, studies on CYP2E1 in fetal and adult tissues were
evaluated (EPA, 1998b). The status of CYP2E1 in human fetuses remains unclear, with
conflicting studies. In those studies showing expression of CYP2E1, levels lower than those in
adults were found (Vieira et al.,. 1996; Boutelet-Bochan et al., 1997; Carpenter et al., 1996;
Hakkola et al., 1998). Regardless of fetal CYP2E1 expression, the enzyme is rapidly induced
upon birth. 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 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 and reported a rapid transcriptional
induction of CYP2E1 (P450) within 1 week following birth that remained elevated throughout 12
weeks. The authors did not quantitate the western blots. However, in this same study, enzyme
activity gradually increased over time, reaching a maximum at adulthood.
EPA OW/OST/HECD 15 Health Risks to Fetuses, Infants, and Children
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I'eno 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 verified, 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 the enzyme activity. On the other hand,
under saturating doses of chloroform, all the available enzyme would be active; thus a 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 gradual increase of CYP2E1 activity
throughout childhood with a maximum level at adulthood, as demonstrated by Vieira et al.
(1996). Therefore, although children may have 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 animal data from
Schenkman et al. (1989) fall within the UF for intraspecies variability.
The next issue to examine is whether the developing fetus may be more susceptible to the
toxicity of chloroform because of its greater rate of cell proliferation. There are very few data for
pre- 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 period of exposure
includes 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 not an increased susceptibility to chloroform that results from pre-
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).
2.2.1.4. Children's Risk in Relation to the MCLG
Because a substantial database indicates that tumor development for chloroform is
secondary to organ toxicity and regenerative proliferation, a nonlinear dose-response approach is
viewed as more appropriate. Use of a low dose-linear approach is considered overly
EPA OW/OST/HECD ~~ 16 Health Risks to Fetuses, Infants, and Children
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conservative for extrapolating cancer risk associated with chloroform exposure (EPA. 1998a,b;
ILSI. 1997). In the nonlinear approach, the MCLG is based on liver toxicity as the most
sensitive effect for chloroform (as it is the lowest possible MCLG 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 susceptible than adults to liver toxicity. The mode of action by which
chloroform produces organ toxicity and carcinogenicity is, thus, the-same for children and adults;
An oral study in dogs (Heywood et al., 1979) was used to derive the R£D of 0.01 mg/kg/day
(EPA, 1994b). This RfD is based on a LOAEL (12.9 mg/kg/day) for hepatotoxicity and
application of a UF of 1,000 (100 was used to account for inter- and intraspecies differences and
a factor of 10 for use of a LOAEL) to calculate the RfD (0.01 mg/kg/day).
The MCLG is calculated to be 0.07 mg/L by assuming an adult tap water consumption of
2 L per day for a 70 kg adult, and by applying a relative source contribution of 20%.
The MCLG of 0.07 mg/L is considered protective of both adults and children, given that
developmental effects occurred at doses above those causing hepatotoxicity. Also, the mode of
action data indicates that children are not uniquely sensitive to the organ toxicity caused by high
doses of chloroform.
Under the linear approach, the MCLG would be zero because under that approach it is
assumed that there is no dose that is without some risk. An MCLG of zero would be a zero risk
for either children or adults.
2.2.2. Brominated Trihalomethanes
There is sufficient evidence for carcinogenicity via ingestion of bromoform and BDCM
to consider them probable human carcinogens. The evidence is limited for DBCM. Based on
available data, mechanism of action involving mutagenicity was postulated for the brominated.
THMs, implying linear low-dose extrapolation as a reasonable approach. The proposed
mechanism of carcinogenicity for these compounds was examined to determine if this would
provide any reason 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. However, an increased risk of this type would be true for all genotoxic carcinogens and
not'
EPA OW/OST/HECD 17 Health Risks to Fetuses, Infants, and Children
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specific to brommated 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.
Brominated THMs are extensively metabolized via oxidative and reductive pathways in
humans and animals, primarily in the liver, but also in the kidney (EPA, 1994b). Oxidative
metabolism (which requires the presence of oxygen) results in the production of a dihalocarbonyl
(CX2O) intermediate. Under conditions of low oxygen (reductive metabolism), the metabolic
.reaction products appear to be free radical species such as the dihalotnethyl radical (CHX2.)-
Both dihalocarbonyls and dihalomethyl radicals are reactive species and may cause
direct/indirect damage .to cellular components including DNA. For BDCM, the genotoxicity is
associated with the glutathione conjugation pathway (Pegram et al., 1997).
Studies have investigated CYP isozyme involvement in THM metabolism. Thorton-
Manning et al. (1994) found that BDCM dosing decreased the activity of CYP 1A and CYP2B,
but not CYP2E1. The cytochrome P450 CYP2E1 subfamily of enzymes undergo changes during
development. Activity is absent during gestation, with onset at birth in rats. Limited studies
compare the enzyme activity at different stages of life for CYP2E1 (Song et al., 1986; Umeno et
al., 1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990; Ronis et al., 1996; Vieira et al.,
1996). There are even fewer data for other P450 isozyme families and glutathione transferase in
this aspect.
Several epidemiology studies have reported an association between exposure to THMs
and developmental/reproductive effects (Bove et al., 1995; Savitz et al., 1995; Nuckols et al.,
1995; Waller et al., 1998). Two studies have reported a relationship specifically between
developmental/reproductive (oxicity and BDCM. An epidemiological study (Kramer et al.,
1992) reported an increased risk of intrauterine growth retardation with exposure to drinking
water with BDCM concentrations greater than or equal to 10 ug/L compared with drinking water
with undetectable BDCM concentrations (OR = 1.7, 95% CI = 0.9-2.9). Waller et al. (1998)
reported an increased risk of spontaneous abortion with consumption of greater than or equal to
five glasses of cold water with a BDCM concentration greater than or equal to 18 ug/L (OR =
2.0, 95% CI = 1.2-3.5). 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.
EPA OW/OST/HECD 18 Health Risks to Fetuses, Infants, and Children
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2.2.2.1. Bromodichloromethane
Developmental/Reproductive Effects
Developmental arid reproductive toxicity data are available for BDCM and were
considered in the derivation of the MCLG, Two oral studies evaluated the developmental
toxicity of BDCM (Ruddick et al., 1983; Narotsky et al., 1997). Ruddick et al. investigated
developmental toxicity in pregnant Sprague-Dawley rats (9-14/group) administered BDCM by
gavage in com oil at dose levels of 0, 50, 100, or 200 mg/kg/day from gestation days 6-15.
Maternal toxicity at the high dose was indicated by a significantly decreased body weight gain.
A NOAEL for developmental effects (sternebra aberration) in this study was 50 mg/kg/day. To
determine the effect of vehicle on BDCM toxicity, Narotsky et al. administered BDCM by
gavage in either com 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 com oil vehicle (83%)
compared with the aqueous vehicle (21%) at the high dose. Accordingly, the NOAEL for
developmental and maternal effects would be 25 mg/kg/day.
Systemic Effects
BDCM causes decreased weight gain and various adverse effects in the nervous and
immune systems, thyroid, kidney, and liver. The NTP (1987) chronic study was selected by EPA
as the most appropriate study for derivation of the RiD. The lowest LOAEL value of 25
mg/kg/day for lesions (in liver, thyroid, and kidney) observed in treated mice was selected to
derive a RfD value of 0.02 mg/kg/day or 0.7 mg/L for a 70 kg adult drinking 2 L of water/day. A
UF of 1,000 was employed for use of a LOAEL, extrapolation from an animal study to humans,
and to account for variation in sensitivity among members of the human population.
Carcinogenicity
There is sufficient animal evidence to consider BDCM a possible human carcinogen by
ingestion. 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
com oil in a 2-year bioassay (NTP, 1987). Consideration of mode of action involving
irreversible changes (mutations) led to use of a linear low-dose extrapolation. The Agency
calculated a cancer oral slope factor based on renal tumors in treated male mice (6.2 x 10"2 per
mg/kg/day). The concentration for 10'5 lifetime risk is 6 ug/L.
EPA OW/OST/HECD 19 Health Risks to Fetuses, Infants, and Children
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Children 's Risk in Relation to the MCLG
The Agency has proposed an MCLG of zero for BDCM based on its probable
carcinogenicity. The consideration of mode of action involving irreversible changes (mutations)
led to the subsequent use of a linear low-dose extrapolation. The Agency believes that the
proposed MCLG of zero is protective of both child and adult health. .
2.2.2.2. Dibromochloromethane
Developmental/Reproductive Effects
Ruddick et al. (1983) investigated developmental toxicity in pregnant Sprague-Dawley
rats (10-12/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. Therefore, the NOAEL in this study for developmental effects
would be 200 mg/kg/day, which is approximately 10-fold higher than the NOAEL used to derive
the RfD based on liver toxicity. 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 0,0.1, 1.0, or 4.0 mg/mL.
DBCM (0, 17, 171, or 685 mg/kg/day). Based on maternal toxicity (weight loss, liver pathology)
and possible fetotoxicity (decreased pup weight and viability in some generations), this study
identified a NOAEL of 17 mg/kg/day and a LOAEL of 171 mg/kg/day for DBCM. This
NOAEL is similar to the NOAEL used to derive the RfD based on liver toxicity (a duration-
adjusted NOAEL of 21 mg/kg/day with a duration-adjusted LOAEL of 43 mg/kg/day).
NTP (1996) conducted a short-term reproductive toxicity study on Sprague-Dawley male
and female rats. These rats were treated with DBCM hi drinking water at concentrations of 0,
50, 150, or 450 ppm during a study period of 35 days (from gestation day 6 through parturition).
Based on measured water consumption, the authors estimated dose levels for the treated males to
be 4.2, 12.4, and 28.2 mg/kg/day, and for treated females 6.3,17.4, and 46.0 mg/kg/day. The
developmental toxicity of the offspring from these treated rats are compared with those from the
control group. After a through examination, no significant reproductive/developmental toxicity
was observed at any dose level; the NOAEL for reproductive/developmental effects identified in
this study is 28.2 mg/kg/day.
EPA OW/OST/HECD . '20 Health Risks to Fetuses, Infants, and Children
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Systemic Effects
DBCM causes decreased weight gain and various adverse effects in the nervous and
immune systems, kidneys, and liver. The NTP subchronic study in rats was selected by EPA as
the most appropriate basis for derivation of the-RID and DWEL (NTP, 1985). The lowest
NOAEL value of 30 mg/kg/day in rats (for absence of clinical and histological changes) was
selected to derive the RfD value of 0.02 mg/kg/day (see also Children's Risk section).
Carcinogenicity
Evidence is limited for DBCM carcinogenicity via ingestion. Tumors occurred in the
livers of female and male mice when these rodents were treated with BDCM in com oil for 2
years (NTP, 1985). Consideration of mode of action involving irreversible changes (mutations)
led to use of a linear low-dose extrapolation. The Agency used a cancer oral slope factor of 8.4 *
10'2 per mg/kg/day based on liver tumors in treated female mice. The concentration for 10"5
lifetime risk is 4 ug/L.
Children's Risk in Relation to the MCLG
The proposed MCLG of 0.06 mg/L for DBCM is based on noncarcinogenic endpoints
(the RfD) with an additional safety factor to account for possible carcinogenicity. An RfD of
0.02 mg/kg/day was derived from a duration-adjusted NOAEL of 21 mg/kg/day for liver toxicity
in exposed rats from a subchronic study (NTP, 1985) and divided by a UF of 1,000. The UF
accounts for use of a less-than-lifetime study, interspecies extrapolation, and variability among
members of the human population. Ah additional safety factor of 10 for possible carcinogenicity
is used to calculate the MCLG along with an assumed drinking water contribution of 80% of
total exposure to DBCM: MCLG = (30 mg/kg/d x 5/7 x 70 kg x 0.8)/(1,000 x 2 L/d x 10) = 0.06
mg/L. Developmental and reproductive toxicity data are available for DBCM and were
considered in the derivation of the RfD value for BDCM. The Agency believes that the MCLG
of 0.06 mg/L is protective of children's health because no developmental or reproductive effects
have been found to occur below the level of the critical effect (liver toxicity) used to derive the
current RfD.
2.2.2.3. Bromoform
Developmental/Reproductive Effects
Ruddick et al. (1983) investigated developmental toxicity in pregnant Sprague-Dawley
rats (14-15/group) administered bromoform by gavage in com oil at dose levels of 0, 50,100, or
200 mg/kg/day from gestation days 6 to 15. No maternal toxicity was observed at any dose
EPA OW/OST/HECD 21 Health Risks to Fetuses, Infants, and Children
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le\-e!: 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 stemebra aberrations increased
in a dose-related fashion. A NOAEL for developmental effects would be 50 mg/kg/day. The
prenatal and postnatal effects of bromoform on fertility and reproduction was investigated by
NTP (1989a) in Swiss CD-I mice using a continuous reproductive breeding protocol. Mice were
administered bromoform by gavage in com oil for 105 days at dose levels of 0, 50, 100, or 200
mg/kg/day. 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 F0
generation. The effect of bromoform administration was also evaluated in the control and high
dose group of the F, generation. No effect was observed in the standard reproductive endpoints
(mating index, fertility index, litter size, proportion of live pups, sex ratio, or pup body weight).
Furthermore, no effect was observed on any sperm parameters evaluated (density, motility, or
morphology). Therefore, the NOAEL for reproductive effects would be 200 mg/kg/day.
t
Systemic Effects
Bromoform causes decreased weight gain and various adverse effects in the nervous and
immune systems, kidney, and liver. The NTP (1989b) subchronic study was selected by EPA as
the most appropriate basis for derivation of the RfD and DWEL. The lowest NOAEL value of
25 mg/kg/day for absence of clinical and histolbgical effect observed in treated rats was selected
to derive the RfD value of 0.02 mg/kg/day. A UF of 1,000 was based on NAS/OW guidelines
for use of a NOAEL from a less-than-lifetime study, extrapolation from an animal study to
humans, and for variation in sensitivity among members of'the human population.
Carcinogenicity
Evidence is sufficient to consider bromoform a possible human carcinogen via ingestion.
Tumors occurred in the large intestine in female and male rats when these rodents were treated
with bromoform in com oil (NTP, 1989b). Consideration of mode of action involving
irreversible changes (mutations) led to use of a linear low-dose extrapolation. The Agency used
a cancer oral slope factor of 7.9 * 10° per mg/kg/day based on liver rumors in treated female
mice. The concentration for 10"s lifetime risk is 40
EPA OW/OST/HECD 22 Health Risks to Fetuses, Infants, and Children
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Children 's Risk in Relation to the MCLG
The Agency has proposed an MCLG of zero for bromoform based on its probable
carcinogenicity. The consideration of mode of action involving irreversible changes (mutations)
led to the use of a linear low-dose extrapolation. The Agency believes that the proposed MCLG
of zero is protective of both child and adult health.
2.2.3. Haloacetic Acids
An MCL of 0.06 mg/L for a combined total of five haloacetic acids (mono-, di-, and
trichloroacetic acids and mono- and dibromoacetic acids) has been determined by the EPA. The
MCLG for DCA has been set at zero based on its potential human carcinogenicity. The MCLG
for TCA has been set at 0.3 mg/L, based on developmental and possible carcinogenic effects
(EPA, 1994a).
2.2.3.1. Dichloroacetic Acid
The MCLG for DCA has been set at zero based on evidence of potential carcinogenicity
in humans and, consideration of mode of action involving irreversible changes (mutations) (EPA,
1994a). The available animal studies also suggest a potential for developmental toxicity;
however, the doses used in these animal studies were very high.
Reproductive/Developmental Effects
In two studies reported by Smith et al. (1992), pregnant Long-Evans female rats
(approximately 20/dose) received DCA by gavage on gestation days 6-15 doses of 900, 1,400,
1,900, or 2,400 mg/kg/day (first study) or 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 1,400 mg/kg/day and above, body
weight gain was significantly reduced at 140 mg/kg/day and above, 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 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. No malformations were observed in 507 fetuses (39 litters) in the control group. The
authors identified a developmental NOAEL of 14 mg/kg/day and LOAEL of 140 mg/kg/day.
Because maternal toxicity was observed at all doses, a LOAEL of 14 mg/kg/day was identified
EPA OW/OST/HECD 23 , Health Risks to Fetuses, Infants, and Children
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for maternal coxicity. The extent to which the developmental effects were attributable to
maternal toxiciry is unknown.
Epstein et al. (1992) reported findings from a similar 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 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 1,900 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 this dose group. In the groups dosed on days
9-11 and 12-15, the mean percentage of cardiac malformations per litter was significantly
(ps0.001) increased.
In the second phase of the study, pregnant dams were administered a single dose of 2,400
mg/kg on gestation days 10, 11, 12, or 13. Fetal weights in each exposure group were similar to
control values. Significant (psO.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 3,500 mg/kg was
administered to dams on gestation days 9,10, 11, 12, or 13. This higher dose level resulted in a
slightly higher incidence of cardiac defects (2.9-3.6%), and the increase was significant (p^O.05)
on day 12. These experiments suggest that increasing the dose has a minimal effect on incidence
of cardiac defects.
Saillenfait et al.'(1995) studied the potential developmental toxicity of DCA in a rat
whole embryo culture system. Groups of 10 explanted embryos from Sprague-Dawley rats were
cultured for 46 hours in 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. 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 5.9 mM and above (Hunter et al., 1996). These in
vitro studies cannot be used for determination of a LOAEL or NQAEL; effects observed in these
experiments cannot be translated to in vivo effects due to pharmacokinetic and
pharmacodynamic uncertainties. However, these results do support the in vivo observations.
EPA OW/OST/HECD 24 Health Risks to Fetuses, Infants, and Children
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Systemic Effects
In a 90-day study in beagle dogs administered DC A by capsule, Katz et al. (1981)
observed effects on body weight gain, hematology, and clinical chemistry parameters at doses of
50, 75, and 100 mg/kg/day; neurotoxic and ocular effects were also observed. In a study by
Cicmanec et al. (1991), in which dogs were administered 12.5-72 mg/kg/day DC A by capsule
for 90 days, a LOAEL of 12.5 mg/kg/day was identified based on degeneration of germinal
epithelium and syncytial giant cell formation in testes, and vacuolization of white myelinated
tracts of cerebrum .and cerebellum. The RfD was calculated from this LOAEL of 12.5 mg/kg/day
using the UF of 3,000 for use of a study with a less-than-lifetime duration, for extrapolation from
an animal study to humans, and to account for variation in sensitivity among members of the
human population.
Carcinogenicity
Several lifetime drinking water studies indicate that DC A is carcinogenic in both rats and
mice (Bull et al., 1990; Daniel et al., 1991; DeAngelo et al., 1991), these studies indicate that
DC A induces liver tumors. Based on these results, the Agency has classified DCA in Group B2:
probable human carcinogen. DCA has been found to be mutagenic and clastogenic, but
responses generally occur at high dose levels. It appears that a mutagenic mechanism for DCA
carcinogenicity may not be important because of the low exposure levels likely to be found in
drinking water. Evidence is still accumulating that suggests a mode of carcinogenic action
through modification of cell signaling systems, with down-regulation of control mechanisms in
normal cells providing a growth advantage to initiated cells (EPA, 1998c). EPA considers that a
contribution of cytotoxicity and compensatory cell proliferation at high doses cannot be ruled out
either. It appears that the shape of tumor dose-response curves for DCA is nonlinear. Currently,
there is an insufficient database for understanding the mode of action or the appropriate basis for
low-dose extrapolation. At this point, EPA has chosen to use an MCLG of zero for DCA, as was
proposed in 1994 (EPA, 1994a). NTP is conducting a 2-year rodent bioassay that will include
full histopathology, and additional mode of action studies are being done by various
investigators, including those at the EPA National Health Effects Research Laboratory.
Children's Risk in Relation to the MCLG
The Agency has proposed an MCLG of zero for DCA based on the considerations
described above. The Agency believes that the proposed MCLG of zero is protective of both
EPA OW/OST/HECD 25 Health Risks to Fetuses, Infants, and Children
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children and adults. Developmental effects were found at higher doses and thus are of secondary
concern compared with the carcinogenic effects of DCA.
2.2.3.2. Trichloro acetic Acid
The MCLG for TCA has been set at 0.3 mg/L based on developmental and possible
carcinogenic effects (EPA, 1994a).
Developmental/Reproductive Effects
In a study by Smith et al. (1989), pregnant Long-Evans rats (20 animals/dose) received
TCA at doses of 0, 330, 800, 1,200, or 1,800 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 dose levels 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 (primarily a defect between the ascending aorta and right
ventricle) occurred in 0%, 32%, 71%, 71%, and 88% of the litters in the 0-, 330-, 800-, 1,200-,
and 1,800 mg/kg/day groups, respectively. The lowest dose, 330 mg/kg/day, was considered the
LOAEL for 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).
Saillenfait et al. (1995) and Hunter et al. (1996) studied developmental toxicity of TCA in
rat whole embryo or mouse whole embryo culture systems. Groups of 10 explanted embryos
from Sprague-Dawley rats were cultured for 24 hrs and groups of 10 from CD-I mice for 46 hrs
in 0-6 mM DCA. A significant dose-related decrease in yolk sac diameters was seen at 1.0 mM
and above, and a significant dose-related decrease was seen in crown-rump length, head length,
somite (embryonic segments) number, protein content, and DNA content at 2.5 mM and above in
rats and at 1.0 mM and above in mice. These in vitro studies cannot be used for determination of
a LO AEL or NOAEL; effects observed in these experiments cannot be translated to in vivo
effects because of pharmacokinetic and pharmacodynamic uncertainties. However, these results
do support the in vivo observations. .
Systemic Effects
In a 90-day study by Mather et al. (1990), Sprague-Dawley rats (10 males/group)
received TCA in'drinking water at dose levels of 0, 50, 500, or 5,000 ppm (0,4.1, 36.5, or 355
EPA OW/OST/HECD 26 Health Risks to Fetuses, Infants, and Children
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mg kg dayi. No effects were seen on body weight or liver and kidney weights at all doses. At
355 mg'kg.day, spleen weight was reduced and the relative kidney and liver weights were
increased. Also, at the high dose, hepatic peroxisomal beta-oxidation activity was increased, but
no effects were seen on hepatic microsomal enzyme activuy. The NOAEL from this study is
36.5 mg/kg/day, and the LOAEL is 355 mg/kg/day.
Bull et al. (1990) treated groups of mice with TCA in their drinking water at 0 mg/L or 1
g/L for 52 weeks, at 2 g/L for 37 weeks with a 15-week recovery period, or at 2g/L for 52 weeks.
Doses calculated were approximately 164 or 329 mg/kg/day for 52 weeks or 309 mg/kg/day for
37 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 is 164 mg/kg/day.
Carcinogenicity
A number of studies have been done to test the carcinogenic potential of TCA, with
conflicting results. A study by DeAngelo et al. (1993) treated groups of 50 male Fischer 344 rats
with 0, 0.05, .0.5, and 5 g/L TCA in drinking water for 104 weeks (0, 2.83, 26, and 283.6
mg/kg/day). Body weight and body weight gain were significantly reduced in the high-dose
group. Decrease of 10% in absolute liver weight was reported in the high-dose group compared
with controls. Histopathology revealed an increase in cytoplasmic necrosis in the high-dose
group but no evidence of hyperplastic nodules in the livers. There was no evidence of
carcinogenicity in any treatment group compared with controls.
An earlier study by DeAngelo and colleagues (1991) in B6C3F, mice showed
hyperplastic nodules and hepatocellular rumors, both adenomas and carcinomas, mostly in male
mice. 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 B6C3F, mice; female mice, however, did not
show these effects.
The Agency has classified TCA in category C—a possible human carcinogen based on
limited evidence of carcinogenicity from experimental studies. As discussed in the 1997 DBF
NODA (EPA, 1997), there have also been several recent studies examining the mode of
carcinogenic action for TCA. These new studies suggest that TCA does not operate through
mutagenic mechanisms. At this time, EPA has determined that the data are not sufficient to
support a choice of mode of action or appropriate basis for a.low-dose extrapolation.
EPA OW/OST/HECD 27 Health Risks to Fetuses, Infants, and Children
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Children 's Risk in Relation to the MCLG
The Agency has proposed an MCLG of 0.3 mg/L for TCA based on a developmental
study in animals. Because this MCLG is based on a NOAEL derived from a
reproductive/development study in animals, the Agency believes that the MCLG of 0.3 mg/L is
protective of children and adults. This number was derived with a consideration of the increased
potential risk to the developing fetus, and is considered to be protective of this subpopulation.
2.2.4. Bromate
Bromate 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.
2.2.4.1. Reproductive/Developmental Effects
Potassium bromate has been studied for prenatal and postnatal reproductive effects in-
multigeneration reproductive studies on rats andmice (EPA, 1993). In a study by Kurokawa et
al. (1990) (cited in EPA, 1993, Final Draft Drinking Water Criteria Document), mice and rats
were fed flour treated with 15 ppm potassium bromate (15 mg/kg/day in diet) over five and eight
generations, respectively. No effects were observed on weight gain, reproductive performance,
or survival in mice or rats. It appeared that 15 mg/kg/day in diet was the NOAEL. A literature
search from 1993 to 1998 did not reveal any new developmental toxicity data for bromate.
2.2.4.2. Systemic Toxicity
The available data are considered insufficient to ascertain noncancer toxic effects of
bromate. Only one toxicity study not dealing with carcinogeincity was located in the literature.
The study failed to provide dose response data and did not identity a NOAEL (EPA, 1993).
2.2.4.3. Carcinogenicity
In the 1994 proposal, EPA concluded that bromate was a probable human carcinogen
(Group B2) under the 1986 EPA Guidelines for Carcinogen Risk Assessment weight of evidence
classification approach, and hence MCLG was set at zero. Cancer is the critical effect. The new
rodent cancer study by DeAngelo et al. (1998) contributes to the weight of evidence for the '
potential human Carcinogenicity of bromate and confirms the study by Kurokawa et al.
(1986a,b). Under the principles of 1996 EPA Proposed Guidelines for Carcinogen Risk
Assessment weight of evidence approach, bromate is considered to be a likely human carcinogen.
This weight of evidence conclusion is based on sufficient experimental findings that include the
EPA OW/OST/HECD 2fjHealth Risks to Fetuses, Infants, and Children
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following: tumors at multiple sites in rats, tumor responses in both sexes, and evidence of
mutagenicity including point mutations and chromosomal aberrations in vitro (EPA, 1998d).
Consideration of mode of action involving irreversible changes (mutations) led to use of a
linear low-dose extrapolation. Cancer risk estimates were derived from the DeAngelo et al.
(1998) study by applying the 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 O.-S g/L, respectively, for 108
weeks. The upper bound cancer potency for bromate ion is estimated to be 0.7 per mg/kg/day-1.
Assuming a daily water consumption of 2 L for a 70 kg adult, lifetime risks of 10", 10'5, and 10"6
are associated with bromate concentrations in water of 5, 0.5, and 0.05 ng/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 presented in the 1994 proposed rule.
2.2.4.4. Children's Risk in Relation to the MCLG
The Agency has proposed an MCLG of zero for bromate based on its probable
carcinogenicity. Consideration of mode of action involving irreversible changes (mutations) led
to use of a linear low-dose extrapolation. The Agency believes that the proposed MCLG of zero
is protective of both children and adults.
2.2.5. Chlorite/Chlorine Dioxide
Chlorite and chlorine dioxide are evaluated together in this assessment for children's risk
because the studies conducted with chlorite, the predominant degradation product of chlorine
dioxide, are likely relevant to characterizing the toxicity of chlorine dioxide. In addition, studies
conducted with chlorine dioxide may be relevant to characterizing the toxicity of chlorite.
Chlorine dioxide is fairly unstable and rapidly dissociates, predominantly into chlorite and
chloride, and to 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 their reaction products with the gastrointestinal
contents. As a result, the toxicity data for one compound are considered applicable for assessing
toxicity for the other.
EPA OW/OST/HECD 29 Health Risks to Fetuses, Infants, and Children
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2.2.5.1. Reproductive/Developmental Effects
EPA (1994c) and the Report of Health Risk Assessment/Characterization of the Drinking
Water Disinfection Byproducts Chlorine Dioxide and Chlorite (EPA, 1998e) describe at length
the reproductive and developmental effects observed in chlorine dioxide and chlorite studies; the
reader is directed to these documents for detailed information. Several developmental toxicity
studies are available on chlorite and chlorine dioxide in addition to the recent two-generation
reproductive rat study designed to evaluate the effects of chlorite onreproduction and pre- and
postnatal development when administered orally via drinking water for two successive
generations (CMA, 1997). The NOAEL from the two-generation reproductive study was
determined to be 35 ppm (3 mg/kg/day) based on neurodevelopmental effects. The data
considered to support this NOAEL are summarized in EPA (1998e) and included the CMA study
as well as previous reports on developmental toxicity. For both chlorite and chlorine dioxide, a
NOAEL dose of 3 mg/kg/day from the CMA two-generation study in rats was considered for the
derivation of an RfD. For chlorite, using a NOAEL of 3 mg/kg/day and applying a UF of 100 to
account for inter- and intraspecies variation in response to toxicity, the MCLG is calculated to be
0.8rhg/L. Using a NOAEL of 3 mg/kg/day from the CMA study and applying a UF of 100 for
inter- and intraspecies variation response to toxicity, the revised MRDLG for chlorine dioxide is
calculated to be 0.8 mg/L.
2.2.5.2. Systemic Toxicity
EPA (1994c) and the Report of Health Risk Assessment/Characterization of the Drinking
Water Disinfection Byproducts Chlorine Dioxide and Chlorite (EPA, 1998e) discuss at length the
subchronic and chronic toxic effects observed in chlorine dioxide and chlorite studies utilizing
various exposure durations; the reader is directed to these documents for detailed information
about these studies. In general, the systemic toxicity studies presented in these documents are of
limited quality. '
2.2.5.3. Carcinogenicity
There have been no long-term oral bioassays for carcinogenicity of chlorine dioxide, and
long-term studies in rats and mice do not provide sufficient evidence to support conclusions as to
the carcinogenic potential of chlorite. In accordance with the 1986 cancer guidelines, chlorine
dioxide and chlorite were categorized in Group D, "Not classifiable as to human carcinogenicity"
(EPA, 1986). In accordance with the 1996 proposed cancer guidelines, the carcinogenicity of
these chemicals is considered, "Cannot be determined" (EPA, 1996).
EPA OW/OST/HECD 30 Health Risks to Fetuses, Infants, and Children
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2.2.5.4. Children 's Risk in Relation to the MCLG
The MCLG and MRDLG calculated for chlorite and chlorine dioxide are considered to be
protective of susceptible groups, including children, given that the RfD is based on a NOAEL
derived from developmental testing, which includes a two-generation reproductive study. In the
case of chlorite and chlorine dioxide a factor .of 10 was used to account for variability between
the average human response and the response of more sensitive individuals including deficiency
associated with glucose-6-phosphate dehydrogenase.
2.2.6. Chlorine
Chlorine forms elemental chlorine (C12), chloride ion (Cl~), and hypochlorous acid
(HOC1) in pure water. As pH increases, hypochlorous acid dissociates to hypochlorite ion
(OC1"). 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).
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, the possibility of oral exposure to chlorine by fetuses,
infants and children is very limited.
2.2.6.1. Developmental/Reproductive Effects
Animal studies have demonstrated no evidence of developmental effects associated with
chlorine (EPA, 1994d). In a study by Carltoh et al. (1986), 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 Utter survival, litter size, and pup weight. The NOAEL in this study
is 5 mg/kg/day; however, higher doses were not tested (IRIS, 1994).
In a multigenerational study, rats were given drinking water chlorinated to a
EPA OW7OST/HECD 31 Health Risks to Fetuses, Infants, and Children
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concentration of 1'JO 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
chlonne, hypochlorous acid, and hypochlorite ion that collectives jccur 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.
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 C12 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,
4rO, or 1.6 mg/kg bw/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 bw/day.
These results were reproduced in retrials of the experiment. No dose of HOC1 was associated
with increases in sperm-head abnormalities (EPA, 1994d).
Six Sprague-Dawley rats (Abdel-Rahman et al., 1982) were administered 0, 1, 10, or 100
mg HOC1/L in drinking water for 2.5 months prior to mating. Animals were maintained on the
treated water after pregnancy was confirmed (day 0) and killed on day 20. Maternal weight at
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. Neither was there a significant difference in numbers of resorptions
between control and treated groups. 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
EPA OW7OST/HECD 32 Health Risks to Fetuses, Infants, and Children
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\Mth 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.
2.2.6.2. Systemic Toxicity
In the National Primary Drinking Water Regulations; Disinfectants and Disinfection
Byproducts; Proposed Rule, issued on July 29, 1994, the MRDLG and the MRDL of 4.0 mg/L
for chlorine was proposed. The study selected for determining a RfD is the 2-year rodent
bioassay that was conducted by the National Toxicology Program (NTP, 1990). In this study,
male and female F344 rats and B6C3F1 mice were given chlorine in distilled drinking water at
levels of 0, 70, 140, and 275 mg/L (0,4, 8, and 14 mg/kg/day) for 2 years. No effects on body
weight or survival were observed for any of the treated groups of animals; it should be noted that
dosing begins when rats and mice are as young as 7 weeks old.
Using a NOAEL of 14 mg/kg/day identified from female rats in the NTP (1990) study, an
MRDLG of 4 mg/L based on lack of systemic toxicity was derived. A UF of 100 was applied to
account for inter- and intraspecies extrapolation in accordance with EPA guidelines when a
NOAEL from a chronic animal study is the basis for the RfD (EPA, 1994d). Given the UF that
is factored into the estimation of the MRDLG and MRDL for chlorine, the value of 4.0 mg/L is
considered protective of sensitive subpopulations.
2.2.6.3. Carcinogenicity
No apparent carcinogenic potential was demonstrated following oral exposure to chlorine
in distilled drinking water as hypochlorite, at levels up to 275 mg/L over a 2-year period (NTP,
1990). The EPA has categorized chlorine in Group D, not classifiable as to human
carcinogenicity (EPA, 1994a).
2.2.6.4. Children's Risk Relative to the MRDLG
The Agency believes that the proposed MRDLG of 4 mg/L is protective of children's
health.
2.2.7. Chloramines
Inorganic chloramines are alternative disinfectants that are rapidly formed when free
chlorine is added to water containing ammonia. Monochloramine is the principal chloramine
formed hi chlorinated natural and wastewaters at neutral pH and is much more persistent in the
environment (EPA, 1994e).
EPA OW/OST/HECD 33 Health Risks to Fetuses, Infants, and Children
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2.2.7.1. Developmental 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. The NOAEL identified in
this study is 100 mg/L (-10 mg/kg/day), which is slightly higher or almost the same as the
NOAEL identified in the systemic toxicity study (see below).
2.2.7.2. Systemic Toxicity
EPA selected the lifetime study in rats (NTP, 1990) as the basis for calculating the
MRDLG for chloramines. F344/N rats were administered 0, 50, 100, and 200 ppm (2.8, 5.3, and
9.5 mg/kg/day) chloramine in drinking water. Although at the highest dose tested (9.5
mg/kg/day) there were statistically significant changes in body and several organ weights, the
biological significance of these changes is unclear. The test animals' consumed a reduced amount
of water, which was perhaps due to unpalatability, and NTP does not consider these changes in
body weight biologically significant. The NOAEL identified is this study is 9.5 mg/kg/day. An
MRDLG of 3 mg/L for chloramine (4 mg/L measured as total chlorine) was 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%. A UF of 100 was applied to the NOAEL
of 9.5 mg/kg/day (10 for intra- and 10 for interspecies variation).
2.2.7.3. Carcinogenicity
The EPA has categorized monochloramine in Group D, not classifiable, because of
inadequate human and animal evidence (EPA, 1994a).
2.2.7.4. Children's Risk in Relation to the MRDLG
The Agency believes that the MRDLG of 3 mg/L is protective of children's health
because no developmental effects were found to occur below the systemic NOAEL of 9.5
mg/kg/day used to derive the current RfD.
EPA OW/OST/HECD 34 Health Risks to Fetuses, Infants, and Children
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3. SUMMARY AND CONCLUSIONS
In developing the Final Stage 1 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:
• Do D/DBPs cause developmental and reproductive effects at doses below those causing
systemic toxicity or cancer?
• Are fetuses and children more susceptible than adults to cancer from D/DBPs?
• Are fetuses and children more susceptible than adults to the systemic toxicity of D/DBPs?
This document evaluates the available data for each of the D/DBPs used for deriving the
MCLG or MRDLG, to determine if the derived MCLGs and/or MRDLGs are protective for the
fetuses and children. Table 3 summarized the comparison of toxicity endpoints for the various
D/DBPs. As can be seen in the table, chloroform, BDCM, bromoform, DCA, and bromate are
considered to be probable human carcinogens. MCLGs of zero were selected after consideration
of the potential carcinogenicity of the chemicals. This MCLG of zero would protect both
children and adults. The MCLG/MRDLGs for DBCM, and chloramine were based on systemic
toxieity. The NOAEL/LOAEL used to derive the numbers are lower than the NOEL/LOAELs
for developmental effects; therefore, the MCLG/MRDLG would be protective of infants and
children. In the case of chlorine, the MCLG 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, chlorite, and TCA, the MCLG/MRDLGs are calculated on
data from developmental studies; hence the numbers derived would be protective for both
children and adults. It can be concluded that the MCLG/MRDLGs of all the D/DBPs in the
Stage 1 D/DBP Rule are protective of fetuses, infants, and children.
EPA OW/OST/HECD 35 Health Risks to Fetuses, Infants, and Children
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