United States Office of Science
Environmental Protection and Technology
Agencf Washington, D.C.
June 30, 2002
EPA-822-R-03-018
DRINKING WATER CRITERIA DOCUMENT ON
BROMINATED TRIHALOMETHANES
DRAFT
February 20, 2003
Prepared For
Health and Ecological Criteria Division
Office of Science and Technology
Office of Water
U.S. Environmental Protection Agency
Washington, D.C. 20460
Under
EPA Contract No. 68-C-99-206
Work assignment 0-16
by
Syracuse Research Corporation
6225 Running Ridge Road
North Syracuse, NY 13212
Under subcontract to
The Cadmus Group, Inc.
135 Beaver Street
Waltham, MA 02452
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FOREWORD
Section 1412 (b) (3) (A) of the Safe Drinking Water Act, as amended in 1986 requires the
Administrator of the Environmental Protection Agency to publish Maximum Contaminant Level
Goals (MCLGs) and promulgate National Primary Drinking Water Regulations for each
contaminant, which, in the judgment of the Administrator, may have an adverse effect on public
health and which is known or anticipated to occur in public water systems. The MCLG is
nonenforceable and is set at a level at which no known or anticipated adverse health effects in
humans occur and which allows for an adequate margin of safety. Factors considered in setting
the MCLG include health effects data and sources of exposure other than drinking water.
This document provides the health effects basis to be considered in establishing the
MCLGs for brominated trihalomethanes found in chlorinated drinking water. To achieve this
objective, data on pharmacokinetics, human exposure, acute and chronic toxicity to animals and
humans, epidemiology and mechanisms of toxicity were evaluated. Specific emphasis is placed on
literature data providing dose-response information. Thus, while the literature search and
evaluation performed in support of this document was comprehensive, only the reports considered
most pertinent in the derivation of the MCLGs are cited in this document. The comprehensive
literature search in support of this document includes information published up to July 2001,
however, more recent information may have been added during the review process.
When adequate health effects data exist, Health Advisory values for less than lifetime
exposure (One-day, Ten-day and Longer-term, approximately 10% of an individual's lifetime) are
included in this document. These values are not used in setting the MCLGs, but serve as informal
guidance to municipalities and other organizations when emergency spills or contamination
situations occur.
Geoffrey Grubbs
Director, Office of Science
and Technology
Office of Water
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Acknowledgments
This document is derived and updated/expanded of the Draft for the Drinking Water Criteria
Document on Trihalomethanes (U.S. EPA, 1994) and the Summary of New Health Effects Data
on Drinking Water Disinfectants and Disinfectant Byproduct (D/DBPs) for the Notice of
Availability (NODA) (U.S. EPA, 1997). This document includes an evaluation of literature on
Brominated Trihalomethanes resulting from a full literature search for toxicity data conducted in
July, 2001. In addition, few newer studies identified after the literature search date have been
included as available at the time of document preparation.
Chemical Manager:
Nancy Chiu, Ph.D.
External Peer Reviewers:
Annette L. Bunge, Ph.D. Colorado School of Mines
John Reif, M.Sc. (Med), D.V.M., Colorado State university
Judy Buelke-Sam, M.A. Toxicology Services
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TABLE OF CONTENTS
I. EXECUTIVE SUMMARY I - 1
II. PHYSICAL AND CHEMICAL PROPERTIES II - 1
A. Properties II - 1
B. Summary II - 2
III. TOXICOKINETICS Ill - I
A. Absorption Ill - 1
B. Distribution Ill - 3
C. Metabolism Ill - 5
D. Excretion Ill - 13
E. Bioaccumulation and Retention Ill - 13
F. Summary Ill - 13
IV. HUMAN EXPOSURE IV - 1
A. Occurrence in Drinking Water IV - 1
1. National Surveys IV - 2
2. Other Studies IV - 8
3. Estimates of Tap Water Ingestion Exposure to Brominated
Trihalomethanes IV - 13
B. Exposure from Sources Other Than Drinking Water IV - 17
1. Dietary Intake IV - 17
2. Air Intake IV - 20
3. Concentrations and Exposures Associated with Swimming Pools and Hot
Tubs IV-27
4. Soil Concentrations and Exposure IV - 30
C. Overall Exposure IV - 30
D. Body Burden IV-31
1. Blood IV-31
2. Mother's Milk IV-35
E. Summary IV - 35
V. HEALTH EFFECTS IN ANIMALS V - 1
A. Acute Exposures V-l
1. Bromodichloromethane V-l
2. Dibromochloromethane V-6
3. Bromoform V-8
B. Short-Term Exposures V-8
1. Bromodichloromethane V-14
2. Dibromochloromethane V-21
3. Bromoform V - 24
C. Subchronic Exposure V - 27
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1. Bromodichloromethane V-30
2. Dibromochloromethane V - 31
3. Bromoform V-33
D. Chronic Exposure V - 34
1. Bromodichloromethane V-34
2. Dibromochloromethane V-38
3. Bromoform V-39
E. Reproductive and Developmental Effects V-41
1. Bromodichloromethane V-41
2. Dibromochloromethane V - 56
3. Bromoform V-59
F. Mutagenicity and Genotoxicitv V - 67
1. Bromodichloromethane V - 67
2. Dibromochloromethane V - 74
3. Bromoform V-81
G. Carcinogenicity V - 85
1. Bromodichloromethane V-85
2. Dibromochloromethane V - 92
3. Bromoform V - 95
H. Other Key Health Effects V - 97
1. Immunotoxicity V - 97
2. Hormonal disruption V - 99
3. Structure-Activity Relationships V - 100
I. Summary V - 101
1. Health Effects of Acute and Short Term Exposure of Animals .... V - 101
2. Health Effects of Longer-term Exposure of Animals V - 101
3. Reproductive and Developmental Effects V - 102
4. Mutagenicity and Genotoxicity V - 103
5. Carcinogenicity Studies in Animals V - 103
6. Other Key effects V - 104
VI. HEALTH EFFECTS IN HUMANS VI - 1
A. Clinical Case Studies VI - 1
1. Bromodichloromethane VI - 1
2. Dibromochloromethane VI - 1
3. Bromoform VI - 1
B. Epidemiological Studies VI - 1
1. Bromodichloromethane VI - 7
2. Dibromochloromethane VI - 11
3. Bromoform VI - 15
C. High Risk Populations VI - 15
D. Summary VI - 16
VII. MECHANISM OF TOXICITY VII - 1
A. Role of Metabolism VII - 1
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B. Biochemical Basis of Toxicity VII - 1
C. Mode of Action of Carcinogenesis VII - 2
D. Interactions and Susceptibilities VII - 5
1. Potential Interactions VII - 5
2. Greater Childhood Susceptibility VII - 6
3. Other Potentially Susceptible Populations VII-13
E. Summary VII - 16
VIII. QUANTIFICATION OF TOXICOLOGICAL EFFECTS VIII - 1
A. Bromodichloromethane VIII - 1
1. Noncarcinogenic effects VIII - 1
2. Carcinogenic Effects VIII - 28
B. Dibromochloromethane VIII - 35
1. Noncarcinogenic effects VIII-35
2. Carcinogenic Effects VIII - 49
C. Bromoform VIII - 55
1. Noncarcinogenic effects VIII - 55
2. Carcinogenic Effects VIII -70
D. Summary VIII - 75
IX. REFERENCES IX - 1
APPENDIX A A - 1
APPENDIX B B - 1
APPENDIX C C - 1
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LIST OF FIGURES
Figure III-l Proposed Metabolic Pathways for Brominated Trihalomethanes Ill - 6
Figure V-2 Proposed Routes for GST-Mediated Metabolic Activation of Trihalomethanes
V - 70
LIST OF TABLES
Table 1-1 Summary of Quantification of Toxicological Effects for Brominated Trihalomethanes
1-15
Table II-1 Physical and Chemical Properties of the Brominated Trihalomethanes II - 1
Table III-l Recovery of Label 8 Hours after Oral Administration of 14C-Labeled Brominated
Trihalomethanes to Male Sprague-Dawley Rats or Male B6C3F, Mice III-l
Table III-2 Cumulative Excretion of Label after Oral Administration of 14C-Labeled
Bromodichloromethane to Male F344 Rats Ill - 2
Table III-3 Over view of Tissue Collection for Analysis of Bromodichloromethane in Sprague-
Dawley Rat Tissues and Fluids (CCC, 2000c) Ill - 4
Table IV-1. Brominated Trihalomethane Concentrations Measured in U.S. Public Drinking Water
Systems Serving 100,000 or More Persons IV - 7
Table I V-2 NRWA Brominated Trihalomethane Results for Small Surface Water Plants
IV - 8
Table IV-3 Bromodichloromethane Concentrations in Drinking Water from the U.S. EPA TEAM
Study (ug/I.) IV - 9
Table IV-4 Dibromochloromethane Concentrations in Drinking Water from the U.S. EPA TEAM
Study IV- 10
Table IV-5 Bromoform Concentrations in Drinking Water from the U.S. EPA
TEAM Study IV - 11
Table IV-6 Estimated Drinking Water Exposures to Brominated Trihalomethanes in U.S. Public
Drinking Water Systems Serving More than 100,000 Persons IV - 14
Table IV-7. Estimated Distribution of Drinking Water Exposures to Brominated Trihalomethanes
for Populations in U.S. EPA TEAM Study IV - 16
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Table IV-8. Selected Concentration Data for Individual Brominated Trihalomethanes (ppt) in
Outdoor Air as Summarized in Brodzinsky and Singh (1983) IV - 22
Table IV-9 Mean Bromodichloromethane Concentrations in Blood Following Three Types of
Water Use Events IV - 33
Table IV-10 Median Tap Water Trihalomethane Levels (ppb) in Cobb County and Corpus Christi
Homes, Water Treatment Plants, and Distribution Systems IV - 34
Table IV-11 Between Site Comparison of Median Blood Levels (ppt) and Changes in Blood
Levels (ppt) after Showering IV - 35
Table V-l Summary of LD50 Values for Brominated Trihalomethanes V - 1
Table V-2 Summary of Acute Toxicity Studies for Brominated Trihalomethanes V - 2
Table V-3 Summary of Short Term Toxicity Studies for Brominated Trihalomethanes .... V - 9
Table V-4 Summary of Subchronic Toxicity Studies for Brominated Trihalomethanes ... V - 28
Table V-5 Summary of Chronic Toxicity Studies for Brominated Trihalomethanes V - 35
Table V-6 NTP (1998) Study Design V - 44
Table V-7 Summary of Experiments Conducted by Bielmeier et al. (2001) V - 46
Table V-8 Mean Consumed Doses (mg/kg-day) of Bromodichloromethane in the Range Finding
Study Conducted by CCC (2000c) V - 49
Table V-9 Summary of Reproductive Studies of Brominated Trihalomethanes V-61
Table V-10 Summary of Mutagenicity, Genotoxicity, and Neoplastic Transformation Data for
Bromodichloromethane V - 72
Table V-l 1 Summary of Mutagenicity, Genotoxicity, and Neoplastic Transformation Data for
Dibromochloromethane V-79
Table V-12 Summary of Mutagenicity, Genotoxicity, and Neoplastic Transformation Data for
Bromoform V - 84
Table V-13 Tumor Frequencies in F344/N Rats and B6C3FX Mice Exposed to
Bromodichloromethane in Corn Oil for 2 Years V - 86
Table V-14 Hepatic and Renal Tumors in Male F344/N Rats Administered
Bromodichloromethane in the Drinking Water for Two Years V - 90
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Table V-15 Frequencies of Liver Tumors in B6C3F, Mice Administered Dibromochloromethane
in Corn Oil for 105 Weeks V - 94
Table V-16 Tumor Frequencies in the Large Intestine of F344/N Rats Exposed to Bromoform in
Corn Oil for 2 Years V - 95
Table VI-1 Epidemiological Studies Investigating an Association Between Chlorinated Drinking
Water and Cancer VI - 2
Table VI-2 Epidemiological Studies Investigating an Association Between Chlorinated Drinking
Water and Adverse Pregnancy Outcomes or Altered Menstrual Function VI - 3
Table VI-3 Means and Adjusted Differences in Menstrual Cycle and Follicular Phase Length by
Quartile of Individual and Summed Brominated Trihalomethanes VI - 13
Table VIII-1 Summary of Candidate Studies for Derivation of the One-day HA for
Bromodichloromethane VIII - 4
Table VIII-2 Summary of Candidate Studies for Derivation of the Ten-day HA for
Bromodichloromethane VIII - 10
Table VIII-3 Summary of Candidate Studies for Derivation of the Longer-term HA for
Bromodichloromethane VIII - 17
Table VIII-4 Summary of Candidate Studies for Derivation of the RfD for
Bromodichloromethane VIII - 22
Table VIII-5 Summary of Preliminary BMD Modeling Results for the Bromodichloromethane
RfD VIII - 25
Table VIII-6 Tumor Frequencies in Rats and Mice Exposed to Bromodichloromethane in Corn
Oil for 2 Years VIII - 32
Table VIII-7 Summary of Cancer Risk Estimates for Bromodichloromethane VIII - 34
Table VIII-8 Summary of Candidate Studies for Derivation of the One-day HA for
Dibromochloromethane VIII - 36
Table VIII-9 Summary of Candidate Studies for Derivation of the Ten-day HA for
Dibromochloromethane VIII - 37
Table VIII-10 Summary of Candidate Studies for Derivation of the Longer-term HA for
Dibromochloromethane VIII - 41
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Table VIII-11 Summary of Candidate Studies for Derivation of the RfD for
Dibromochloromethane
VIII - 45
Table VIII-12 Results of Preliminary BMD Modeling of Selected Data from NTP (1985)\Stid-ids7
Table VIII-13 Frequencies of Liver Tumors in Mice Administered Dibromochloromethane in
Corn Oil for 105 Weeks VIII - 53
Table VIII-15 Summary of Candidate Studies for Derivation of the Ten-day HA for Bromoform
VIII - 57
Table VIII-16 Summary of Candidate Studies for Derivation of the Longer-term HA for
Bromoform VIII - 60
Table VIII-17 Summary of Candidate Studies for Derivation of the RfD for Bromoform
VIII - 66
Table VIII-18 Tumor Frequencies in Rats Exposed to Bromoform in Corn Oil for 2 Years
VIII - 73
Table VIII-19 Carcinogenic Risk Estimates for Bromoform VIII - 74
Table VIII-20 Summary of Advisory Values for Bromodichloromethane,
Dibromochloromethane, and Bromoform VIII - 75
Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane A - 3
Table A-2 Candidate Studies and Data for BMD Modeling -Dibromochloromethane .... A - 15
Table A-3 Candidate Studies and Data for BMD Modeling - Bromoform A - 24
Table A-4 Model Equations used in BMD Calculations for Health Advisories A - 29
Table A-5 Benchmark Dose Modeling Results for Bromodichloromethane A - 41
Table A-6 Benchmark Dose Modeling Results for Dibromochloromethane A - 50
Table A-7 Benchmark Dose Modeling Results for Bromoform A - 56
Table C-l DBCM Concentrations Measured in U.S. Public Drinking Water Systems Serving
100,000 or More Persons C-4
Table C-2 Selected Concentration Data for Individual Brominated Trihalomethanes (ppt) in
Outdoor Air as Summarized in Brodzinsky and Singh (1983) C-5
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Table C-3 Results of RSC Calculations for DBCM
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I. EXECUTIVE SUMMARY
Brominated trihalomethanes are volatile organic liquids that have a number of industrial
and chemical uses. The chief reason for health concern is that they are generated as by-products
during the disinfection of drinking water. The brominated trihalomethanes occurring in water are
bromoform, dibromochloromethane, and bromodichloromethane. These compounds are formed
when hypochlorous acid oxidizes any bromide ion present in water to form hypobromous acid,
which subsequently reacts with organic material to form the brominated trihalomethanes.
Toxicokinetics
No human data on absorption of brominated trihalomethanes are available.
Measurements in mice and rats indicate that gastrointestinal absorption of brominated
trihalomethanes is rapid (peak levels attained less than an hour after administration of a gavage
dose) and extensive (63% to 93%). Most studies of brominated trihalomethane absorption have
used oil-based vehicles. A study in rats found that the initial absorption rate of
bromodichloromethane was higher when the compound was administered in an aqueous vehicle
than when administered in a corn oil vehicle.
Data for distribution of brominated trihalomethanes in human organs and tissues are
limited. Bromoform was found primarily in the stomach and lungs of a human overdose victim,
with lower levels detected in intestine, liver, kidney and brain. Dibromochloromethane was found
in 1 of 42 samples of human breast milk collected from women living in urban areas.
Radiolabeled brominated trihalomethanes were detected in a variety of tissues following oral
dosing in rats and mice. Approximately 1 to 4% of the administered dose was recovered in body
tissues when analysis was conducted 8 or 24 hours post-treatment. The highest concentrations
were detected in stomach, liver, blood, and kidneys when assayed 8 hours after administration of
the compounds. Bromodichloromethane was detected at a concentration of 0.38 |ig/g in the milk
of one of three female rats exposed to approximately 112 mg/kg-day during a
reproductive/developmental study. Bromodichloromethane was not detected in placentas,
amniotic fluid, or fetal tissue collected on gestation day 21 from rats exposed to doses up to
approximately 112 mg/kg-day or in plasma collected from postpartum day 29 weanling pups.
Bromodichloromethane was detected at concentrations slightly above the limit of detection in
placentas from two litters born to rabbits exposed to 76 mg/kg-day. Bromodichloromethane was
detected in one fetus from a rabbit exposed to 76 mg/kg-day "...at a level below the limit of
detection". Bromodichloromethane was not detected in placentas from female rabbits exposed to
doses of approximately 32 mg/kg-day, or in amniotic fluid or the remaining fetuses from rabbits
exposed to doses of approximately 76 mg/kg-day.
Brominated trihalomethanes are extensively metabolized by animals. Metabolism of
brominated trihalomethanes occurs via two pathways. One pathway predominates in the presence
of oxygen (the oxidative pathway) and the other predominates under conditions of low oxygen
tension (the reductive pathway). In the presence of oxygen, the initial reaction product is
trihalomethanol (CX3OH), which spontaneously decomposes to yield the corresponding
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dihalocarbonyl (CX20). The dihalocarbonyl species are reactive and may form adducts with
cellular molecules. When intracellular oxygen levels are low, the trihalomethane is metabolized
via the reductive pathway, resulting in a highly reactive dihalomethyl radical (*CHX2), which may
also form covalent adducts with cellular molecules. The metabolism of brominated
trihalomethanes and chloroform appear to occur via the same pathways, although in vitro and in
vivo data suggest that metabolism via the reductive pathway occurs more readily for brominated
trihalomethanes. Both oxidative metabolism and reductive metabolism of trihalomethanes appear
to be mediated by cytochrome P450 isoforms. The identity of cytochrome P450 isoforms that
metabolize brominated trihalomethanes has been investigated in several studies which used
bromodichloromethane as a substrate. The available data suggest that the cytochrome P450
isoforms CYP2E1, CYP2B1/2, and CYP1A2 metabolize bromodichloromethane in rats. The
human isoforms CYP2E1, CYP1A2, and CYP3A4 show substantial activity toward
bromodichloromethane in vitro and low but measurable levels of CYP2A6 activity have also been
detected. Based on the available data, CYP2E1 and CYP1A2 are the only isoforms active in both
rats and humans. CYP2E1 shows the highest affinity for bromodichloromethane in both species
and the metabolic parameters Km and kcal are similar for rat and human CYP2E1. In contrast, the
metabolic parameters for CYP1A2 differ in rats and humans. The pattern of results for isozyme
activity obtained from an inhalation study of bromodichloromethane was similar to the pattern
reported for male F344 rats treated with bromodichloromethane by gavage.
The lung is the principle route of excretion in rats and mice. Studies with 14C-labeled
compounds indicate that up to 88% of the administered dose can be found in exhaled air as
carbon dioxide, carbon monoxide, and parent compound. Excretion in the urine generally appears
to be 5% or less of the administered oral dose. Data from one study suggest that fecal excretion
accounts for less than 3% of the administered dose.
Human Exposure
Brominated trihalomethanes are found in virtually all water treated for drinking; however,
concentrations of individual forms vary widely depending on the type of water treatment, locale,
time of year, sampling point in the distribution system, and source of the drinking water.
Occurrence data for brominated trihalomethanes are available from 13 national surveys and 9
additional studies that are more restricted in scope. The procedures used for sampling processing
and storage and calculation of summary statistics should be carefully considered when evaluating
and comparing brominated trihalomethane occurrence data. Some methods restrict
trihalomethane formation by refrigeration or the use of quenching agents, whereas others
maximize trihalomethane formation by storage at room temperature. Approaches to data
summarization vary in their treatment of data below the analytical detection level or minimum
reporting level.
When all available national survey data are considered, bromodichloromethane concen-
trations in drinking water range from below the detection limit to 183 |ig/L (ppb), while dibromo-
chloromethane and bromoform concentrations range from below the detection limit to 280 |ig/L
(ppb). When data for the three brominated trihalomethanes are compared, the frequency of
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detection and measured concentrations of bromodichloromethane in drinking water supplies tend
to be higher than those for dibromochloromethane. Bromoform is detected less frequently and at
lower concentrations than the other two brominated trihalomethanes, except in some ground
waters. Concentrations of all trihalomethanes in drinking water were generally lower when the
raw water is obtained from ground water sources rather than surface water sources. The most
recent national survey data are those collected by the U.S. EPA under the Information Collection
Rule (ICR). Monitoring data were collected over an 18-month period between July 1997 and
December 1998 from approximately 300 water systems operating 501 plants and serving at least
100,000 people. Summary occurrence data stratified by raw water source (groundwater or
surface water) are available for finished water, the distribution system (DS) average, and the DS
high values. The mean, median, and 90th percentile values for surface water DS average
concentrations in the ICR survey are 8.6, 70.2, and 20.3 |ig/L, respectively, for
bromodichloromethane (range of individual values 0 - 65.8 |ig/L); 2.4, 4.72, and 13.2 |ig/L,
respectively, for dibromochloromethane (range 0 - 67.3); and 0. 1.18, and 3.10, respectively, for
bromoform (range 0-3.43).
Exposure to brominated trihalomethanes via ingestion of drinking water was estimated
using data obtained for disinfectants and disinfection byproducts under the Information Collection
Rule (ICR). ICR data offer several advantages over other national studies for purposes of
estimating national exposure levels of adults in the United States to brominated trihalomethanes
via ingestion of drinking water. First, they are recent and reflect relatively current conditions.
Second, data of very similar quality and quantity were collected systematically from a large
number of plants (501) and systems (approximately 300), including both surface and ground
water systems. Third, the mean, median, and 90th percentile value were estimated on the basis of
all samples taken, not just the sample detects. Thus, these descriptive statistics are representative
of the exposures of the entire populations served by those systems, not just the populations served
by systems with higher concentrations of these compounds. However, this study can not be
considered representative of smaller public water supplies or water supplies from the most highly
industrialized or contaminated areas.
Exposure was calculated by multiplying the concentration of individual brominated
trihalomethanes in drinking water by the average daily intake, assuming that each individual
consumes two liters of water per day. The annual median, mean, and upper 90th percentile values
are presented for both surface and ground water systems. Assuming that the DS High value
actually represents the average exposure level of persons served by one plant distribution pipe
with the longest water-residence time, the DS High value might be used to estimate a high-end
exposure level.
For bromodichloromethane, the median, mean, and 90th percentile population exposures
from surface water systems are estimated to be 17, 20, and 40 |ig/person/day, respectively. The
same values for populations exposed to bromodichloromethane from ground water systems are
lower - 3.6, 8.1, and 22 |ig/person/day, respectively. For dibromochloromethane, the median,
mean, and 90th percentile population exposures from surface water systems are estimated to be
4.8, 9.4, and 26 |ig/person/day, respectively. The corresponding values for populations exposed
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to dibromochloromethane from groundwater system are lower, with estimates of 2.7, 6.2, and 18
|ig/person/day, respectively. For bromoform, the median, mean, and 90th percentile population
exposures from surface water systems are estimated to be near 0, 2.4, and 6.2 |ig/person/day,
respectively. The same values for populations exposed to bromoform from ground water systems
are higher, with estimates of 0.65, 3.8, and 9.6 |ig/person/day, respectively.
For purposes of comparison, estimates of ingestion exposure to bromodichloromethane,
dibromochloromethane, and bromoform in drinking water were also estimated from data collected
in other, older studies. Ingestion from ground water supplies was estimated from the median
levels found in the Ground Water Supply Survey conducted by U.S. EPA in 1980-81. Based on
the range of median levels (1.4-2.1 |ig/L (ppb)) and a consumption rate of two liters per day, the
median ingestion exposure to bromodichloromethane may range from 2.8 to 4.2 |ig/day.
Similarly, median exposure to dibromochloromethane may range from 4.2 to 7.8 |ig/day, and for
bromoform, median exposure may range from 4.8 to 8.4 |ig/day. Exposure to
bromodichloromethane from surface water supplies can be estimated based on the range of
median values observed under different conditions in the National Organics Monitoring Survey
conducted by U.S. EPA in 1976-1977, which mainly sampled surface water systems. Based on a
range of 5.9-14 |ig/L (ppb), exposure to bromodichloromethane from surface water is estimated
to be between 12 and 28 |ig/day. Similarly, based on the range of medians reported for
dibromochloromethane concentrations, the median exposure is estimated to be up to 6 |ig/day.
The median levels of bromoform in the surface water supplies have been found to be less than the
EPA Drinking Water minimum reporting levels (MRLs) of 0.5-1 |ig/L (ppb). An estimate of
exposure based on the MRLs will be overly conservative because the actual concentration of
bromoform is not detectable. Based on the range of MRLs, 0.5-1 |ig/L (ppb), the exposure to
bromoform is estimated to range from 1 to 2 |ig/day for surface water supplies.
Ingestion exposure to brominated trihalomethanes in drinking water can also be estimated
from the concentrations found at the tap in the U.S. EPA's Total Exposure Assessment
Methodology (TEAM) study. Estimates of the average of the population intakes for ingestion of
bromodichloromethane from drinking water range from 0.42 to 42 //g/person/day. The upper 90th
percentile estimates range from <2.0 to 90 /ig/person/day. Estimates of the average population
intake of dibromochloromethane from drinking water range from 0.2 to 56 //g/person/day. The
upper 90th percentile estimates range from < 0.9 to 86 /ig/person/day. Estimates of the average of
the population intakes of bromoform, for those areas in which bromoform was measurable in a
majority of the samples, range from 1.6 to 16.2 //g/person/day. The upper 90th percentile
estimates range from 2.4 to 26 //g/person/day. Four of the six locations in the TEAM study,
however, had a low frequency (less than 10%) of detection of bromoform in measurable
quantities.
Sources of uncertainty in these estimates of ingestion exposure include use of different
analytical methods, failure to report quantitation limits, using measures near the detection limit,
failure to report how nondetects are handled when averaging values (e.g., set to zero or one half
the detection limit), and failure to report sample storage method and duration. In addition, many
environmental factors influence the concentrations of these compounds in drinking water at the
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tap and in vended or bottled waters used for drinking. These factors include season and
temperature, geographic location, source of water, residence time in distribution system, and
others.
Relatively few studies have analyzed non-beverage foods for the occurrence of brominated
trihalomethanes. In the few studies available, bromodichloromethane has been detected in non-
beverage foods (i.e., in one sample of butter at 7 ppb, in three samples of ice-cream at 0.6 to 2.3
ppb, in 6 of 10 samples of bean curd at 1.2 to 5.2 ppb, and in one sample of bacon). In addition,
bromodichloromethane was detected in one sample each of eleven foods out of 70 tested in 14
Market Baskets for the FDA Total Diet Study. The detected concentrations ranged from 10 to
37 ppb for individual food items. Studies that analyzed non-beverage foods for
dibromochloromethane and bromoform detected neither compound in any of the tested samples.
Brominated trihalomethanes have been detected in up to a third or one half of the types of
prepared beverages examined in some studies, being detected most frequently in colas and other
carbonated soft drinks. Bromodichloromethane has been found most frequently of the three
compounds and bromoform the least frequently. Bromodichloromethane was detected in
approximately half of the prepared beverages examined by McNeal et al. (1995) in the United
States and in all of 13 soft drinks that they analyzed. One sampled soft drink contained
bromodichloromethane at a concentration of 12 ppb; the remainder of the samples contained less
than 4 ppb. Bromodichloromethane was detected in one sample of fruit juice at 5 ppb.
Some data on the occurrence of brominated trihalomethanes in foods and beverages are
available from studies conducted in Italy, Japan, and Finland. These studies were also limited in
scope to examination of relatively few food or beverage items. Bromodichloromethane,
dibromochloromethane, and bromoform concentrations measured in foods and beverages in Italy,
Japan and Finland ranged from undetectable to 40 ppb, undetectable to 13.9 ppb, and
undetectable to 10.7 ppb, respectively. Because of possible differences in water disinfection or
food processing practices, these data may not be representative of concentrations in foods and
beverages produced in the U.S.
Measured concentrations of brominated trihalomethanes in outdoor air are variable from
site to site. When data from several urban/suburban and source-dominated sites in Texas,
Louisiana, North Carolina and/or Arkansas were combined, the resulting average outdoor air
concentrations were 110 ppt (0.74 |ig/m3) for bromodichloromethane, 3.8 ppt (0.032 |ig/m3) for
dibromochloromethane, and 3.6 ppt (0.037 |ig/m3) for bromoform. A regional study conducted at
several sites in southern California found bromodichloromethane, dibromochloromethane, and
bromoform in 35%, 17%, and 31% of the samples, respectively. The maximum concentrations
observed were 40 ppt (0.27 |ig/m3) for bromodichloromethane; 290 ppt (2.5 |ig/m3) for
dibromochloromethane; 310 ppt (3.2 |ig/m3) for bromoform. Bromodichloromethane was
detected in 64% (n=l 1) and 17% (n=6) of personal air samples collected in Texas and North
Carolina. The detected concentrations ranged from 0.12 to 4.36 |ig/m3 (0.017 to 0.65 ppb).
Dibromochloromethane was not detected.
Mean concentrations in indoor air ranged from 0.38 to 0.75 |ig/m3 for bromodichloro-
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methane; 0.44 to 0.53 |ig/m3 for dibromochloromethane, and 0.29 to 0.35 |ig/m3 for bromoform,
as determined from 15 minute samples collected in 48 New Jersey residences. It was not clear
whether these values were based exclusively on detected concentrations. In a separate study,
levels of brominated trihalomethanes in indoor air were locally increased (e.g., in shower/bath
enclosures and vanity areas) during showering and bathing events. The levels of individual
brominated trihalomethanes in air were reported to be consistent with the levels in tap water.
No data for occurrence of brominated trihalomethanes in soil were available in the
materials reviewed for this document. The chemical and physical properties of the brominated
trihalomethanes indicate that they should volatilize readily from wet or dry soil surfaces.
Therefore, ingestion of soil is not expected to be a significant route of exposure.
Brominated trihalomethanes have been detected in the blood and breast milk of humans.
The level of individual brominated trihalomethanes in blood increases shortly after exposure to
these compounds in tap water (by dermal contact and/or inhalation of the volatilized compound)
during bathing and showering. Dibromochloromethane was detected in one of eight samples of
breast milk collected from women living in the vicinity of U.S. chemical manufacturing plants or
user facilities.
The RSC (relative source contribution) is the percentage of total daily exposure that is
attributable to tap water when all potential sources are considered (e.g., air, food, soil, and
water). Ideally, the RSC is determined quantitatively using nationwide, central tendency and/or
high-end estimates of exposure from each relevant medium. In the absence of such data, a default
RSC ranging from 20% to 80% may be used.
The RSC used in the current and previous drinking water regulations for
dibromochloromethane is 80%. This value was established by use of a screening level approach
to estimate and compare exposure to dibromochloromethane from various sources. Information
considered for during this process is summarized in Appendix C. There are some uncertainties in
the 80%) RSC that are related to the availability of adequate concentration data for
dibromochloromethane in media other than water. Parallel RSC calculations were not performed
for bromodichloromethane and bromoform. The EPA has set the regulatory level for these
chemicals in drinking water at zero because it has been determined that they are probable human
carcinogens. Therefore, determination of an RSC is not relevant for these chemicals because it is
the Agency's policy to perform RSC analysis only for noncarcinogens.
The use of chlorine to disinfect swimming pools and hot tubs results in the formation of
brominated trihalomethanes. Swimming pool and hot tub users are potentially exposed to
brominated trihalomethanes via dermal contact, ingestion, and inhalation of compounds released
to the overlying air. As a result, swimming pool and hot tub users may experience greater overall
exposures to brominated trihalomethanes than the general population. One study indicated that
bromodichloromethane, dibromochloromethane, and bromoform concentrations in swimming pool
and hot tub water ranged from 1 to 105 |ig/L (ppb), from 0.1 to 48 |ig/L (ppb), and from less
than 0.1 to 62 |ig/L (ppb), respectively. Concentrations of the same brominated trihalomethanes
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in the air two meters above the pool water ranged from less than 0.1 to 14 |ig/m3 (0.015-2.09
ppb), from less than 0.1-10 |ig/m3 (0.011-1.2 ppb), and from less than 0.1 to 5.0 |ig/m3
(0.0097-0.48 ppb), respectively. Data from several studies confirm the uptake of brominated
trihalomethanes from swimming pools, hot tubs, and environs by dermal and/or inhalation
pathways.
Health Effects of Acute and Short-term Exposure of Animals
Large oral doses of brominated trihalomethanes are lethal to mice and rats. Reported
acute LD50 values range from 450 to 969 mg/kg for bromodichloromethane, 800 to 1,200 mg/kg
for dibromochloromethane, and 1,388 to 1,550 mg/kg for bromoform.
Acute oral exposure to sublethal doses of brominated trihalomethanes can also produce
effects on the central nervous system, liver, kidney, and heart. Ataxia, anaesthesia, and/or
sedation were noted in mice receiving 500 mg/kg bromodichloromethane, 500 mg/kg
dibromochloromethane, or 1,000 mg/kg bromoform. Renal tubule degeneration, necrosis, and
elevated levels of urinary markers of renal toxicity have been observed in rats receiving 200 to
400 mg/kg bromodichloromethane. Elevated levels of serum markers for hepatotoxicity and
have been observed in rats at doses of bromodichloromethane ranging from approximately 82 to
400 mg/kg-day, and hepatocellular degeneration and necrosis were observed at 400 mg/kg.
Effects on heart contractility were reported in male rats at doses of 333 and 667 mg/kg
dibromochloromethane.
Short-term oral exposure of laboratory animals to brominated trihalomethanes has been
observed to cause effects on the liver and kidney. Hepatic effects, including organ weight
changes, elevated serum enzyme levels, and histopathological changes, became evident in mice
and/or rats administered 38 to 250 mg/kg-day bromodichloromethane, 147 to 500 mg/kg-day
dibromochloromethane, or 187 to 289 mg/kg-day bromoform for 14 to 30 days. Kidney effects,
characterized by decreased p-aminohippurate uptake, histopathological changes, and organ weight
changes, became evident in mice and/or rats administered 148 to 300 mg/kg-day
bromodichloromethane, 147 to 500 mg/kg-day dibromochloromethane, or 289 mg/kg-day
bromoform for 14 days. Evidence for decreased immune function was noted at
bromodichloromethane and dibromochloromethane doses of 125 mg/kg-day.
The inhalation toxicity of bromodichloromethane has been evaluated in wild type and p53
heterozygous FVB/N and C57BL/N mice. Dose-related renal tubular degeneration, and
associated regenerative cell proliferation were seen in all strains at concentrations of 10 ppm and
above after one week of exposure. Dose-related increases in hepatic degeneration and
regenerative cell proliferation were observed at 30 ppm and above. After three weeks of
exposure, macroscopic and histologic lesions in the kidney and liver were resolved and cell
proliferation indices had returned to near baseline levels. Pathological changes were more severe
in the FVB/N compared to the C57BL/N mice and were more severe in the heterozygotes than in
the wild type strains.
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Health Effects of Subchronic and Chronic Exposure of Animals
The predominant effects of subchronic oral exposure occur in the liver and kidney. The
effects produced in these two organs are similar in nature to those described for short-term
exposures, with liver appearing to be the most sensitive target organ for dibromochloromethane
and bromoform exposure. Histopathological changes in the liver were reported in mice and/or
rats administered 200 mg/kg-day bromodichloromethane, 50 to 250 mg/kg-day dibromochloro-
methane, or 50 to 283 mg/kg-day bromoform. Histopathological changes in the kidney were
reported in mice and/or rats administered 100 mg/kg-day bromodichloromethane, or 250 mg/kg-
day dibromochloromethane.
As observed for shorter durations of exposure, the predominant effects of chronic oral
exposure are observed in the liver and kidney. Histopathological signs of hepatic toxicity in mice
and/or rats became evident at doses of 6 to 50 mg/kg-day for bromodichloromethane, 40 to 50
mg/kg-day for dibromochloromethane, and 90 to 152 mg/kg-day for bromoform. Signs of
bromodichloromethane-induced renal toxicity became evident in mice and rats treated with doses
of 25 and 50 mg/kg-day, respectively.
Reproductive/Developmental Effects in Animals
Reproductive and developmental studies of brominated trihalomethanes are summarized in
Table V-9. Data on the developmental effects of brominated trihalomethanes suggest that these
chemicals are toxic to the fetus in most cases only at doses that result in maternal toxicity. Signs
of maternal toxicity (decreased body weight, body weight gain and/or changes in organ weight)
were reported in rats administered bromodichloromethane at 25 to 200 mg/kg-day and in rabbits
administered 4.9 to 35.6 mg/kg-day. Signs of maternal toxicity were observed in rats or mice
administered 17 (marginal) to 200 mg/kg-day dibromochloromethane and in mice administered
100 mg/kg-day bromoform. Maternal toxicity was not observed in female rats dosed with up to
200 mg/kg-day of bromoform. Several well-conducted studies on the developmental toxicity of
bromodichloromethane gave negative results at doses up to 116 mg/kg-day in rats and 76 mg/kg-
day in rabbits when administered in drinking water. However, in other studies, slightly decreased
numbers of ossification sites in the hindlimb and forelimb were observed in fetuses of Sprague-
Dawley rats administered 45 mg/kg-day in the drinking water on gestation days 6 to 2land
sternebral aberrations were observed in the offspring of Sprague-Dawley rats administered 200
mg/kg-day by gavage in corn oil. Reductions in mean pup weight gain and pup weight were
observed when the pups were administered bromodichloromethane in the drinking water at
concentrations of 150 ppm and above (biologically meaningful estimates of intake on a mg/kg-day
basis could not be calculated for this study). Full litter resorption has been noted in F344 rats, but
not Sprague-Dawley rats, treated with bromodichloromethane at doses of 50 to 100 mg/kg-day
during gestation days 6 to 10. Chronic oral exposure of male F344 rats to bromodichloromethane
resulted in reduced sperm velocities at doses of 39 mg/kg-day. This response was not
accompanied by histopathological changes in any reproductive tissue examined. Adverse clinical
signs and reduced body weight and body weight gain were observed in parental generation female
rats and F, male and female rats at 150 ppm (approximately 11.6 to 40.2 mg/kg-day) in a two
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generation drinking water study of bromodichloromethane. In the same study, small but
statistically significant delays in sexual maturation occurred in F1 males at 50 ppm
(approximately 11.6 to 40.2 mg/kg-day) and F1 females at 450 ppm (approximately 29.5 to 109
mg/kg-day). These delays may have been secondary to dehydration caused by taste aversion to
bromodichloromethane in the drinking water.
Four of five studies on the reproductive or developmental toxicity of dibromochloro-
methane gave negative results when tested at doses of up to 200 mg/kg-day. In the fifth study,
dibromochloromethane administered at 17 mg/kg-day in a multigenerational study resulted in
reduced day 14 postnatal body weight in one of two F2 generation litters. At 171 mg/kg-day, the
mid-dose in the study, decreased litter size, viability index, lactation index, and postnatal body
weight were observed in some F1 and/or F2 generations.
The developmental and reproductive toxicity of bromoform has been examined in two
studies. Bromoform administered to Sprague-Dawley rats at 100 mg/kg-day in corn oil by
gavage resulted in a significant increase in sternebral aberrations in the apparent absence of
maternal toxicity. In a continuous breeding toxicity protocol, gavage doses of 200 mg/kg-day in
corn oil resulted decreased postnatal survival, organ weight changes, and liver histopathology in
F1 ICR Swiss mice of both sexes. No effects on fertility or other reproductive endpoints were
noted.
Mutagenicity and Carcinogenicity Studies
In vitro and in vivo studies of the mutagenic and genotoxic potential of
bromodichloromethane, dibromochloromethane, and bromoform have yielded mixed results.
Synthesis of the overall weight of evidence from these studies is complicated by the use of a
variety of testing protocols, different strains of test organisms, different activating systems,
different dose levels, different exposure methods (gas versus liquid), and in some cases, problems
due to evaporation of the test chemical. Overall, a majority of studies yielded more positive
results for bromoform and bromodichloromethane. The genotoxicity and mutagenicity data for
dibromochloromethane are inconclusive. Recent studies in strains of Salmonella engineered to
contain rat theta-class glutathione S-transferase suggest that mutagenicity of the brominated
trihalomethanes may be mediated by glutathione conjugation.
Carcinogenicity Studies in Animals
The carcinogenic potential of individual brominated trihalomethanes has been investigated
in oral exposure studies conducted in mice and rats. Brominated trihalomethanes administered
individually in drinking water induced the formation of aberrant crypt foci (ACF), putative
preneoplastic lesions, in the colons of male F344/N rats, but not male B6C3F, mice. However, a
cancer bioassay of bromodichloromethane administered in the drinking water did not detect colon
cancer in male F344/N rats. Ingestion of bromodichloromethane in corn oil significantly increased
the incidence of liver tumors in female mice, renal tumors in male mice and in male and female
rats, and tumors of the large intestine in male and female rats. A drinking water exposure study of
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bromodichloromethane in rats reported a significant induction of liver tumors at the lowest dose
tested, but not at higher doses. Administration of dibromochloromethane in corn oil significantly
increased the incidence of liver tumors in male and female mice. Administration of bromoform in
corn oil increased the incidence of uncommon intestinal tumors in male and female rats, with the
effect reaching statistical significance in females. No data are available for the carcinogenicity of
brominated trihalomethanes via the inhalation or dermal routes.
Other Key Health Effects Data from Animal Studies
The immunotoxicity of brominated trihalomethanes has been investigated in mice and rats.
Short-term bromodichloromethane exposure resulted in decreased antibody forming cells in
serum, decreased hemagglutinin titers, and/or suppression of Con A-stimulated proliferation of
spleen cells at doses of 125 to 250 mg/kg-day.
Studies in pregnant F344 rats detected decreased levels of progesterone in animals
administered 75 or 100 mg/kg bromodichloromethane by aqueous gavage on gestation day 8 or 9.
Increased levels of luteinizing hormone were observed two to three days after dose
administration. Disruption of luteal responsiveness to luteinizing hormone has been proposed as a
possible mode of action by which bromodichloromethane elicits full litter resorption in F344 rats,
but additional data are required to confirm this hypothesis. The available evidence does not
suggest that dibromochloromethane or bromoform affect hormonal parameters.
Limited structure-activity data for brominated trihalomethanes and chloroform suggest
that bromination may influence the proportion of compound metabolized via the oxidative and
reductive pathways, with brominated compounds being more extensively metabolized by the
reductive pathway. Additional evidence suggests that a GST-mediated pathway may play an
important role in metabolism of brominated trihalomethanes.
Health Effects in Humans
Limited human health data are available for the brominated trihalomethanes. In the past,
bromoform was used as a sedative for children with whooping cough. Doses of 50 to 100 mg/kg-
day usually produced sedation without apparent adverse effects. Some rare instances of death or
near-death were reported, although these cases were generally attributed to accidental overdoses.
No human toxicological data were available for bromodichloromethane or
dibromochloromethane.
Numerous epidemiological studies have examined the association between water
chlorination and increased cancer mortality rates. None of these studies has identified a strong
association between cancer and exposure to any individual brominated trihalomethane.
Several epidemiological studies have examined the association of chlorinated water use
with various pregnancy outcomes, including low birth weight, premature birth, intrauterine
growth retardation, spontaneous abortion, stillbirth , and birth defects. An association has been
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reported for exposure to bromodichloromethane (or a closely associated compound) and a
moderately increased risk of spontaneous abortion during the first trimester. A confirmation of
this finding is pending reanalysis of the original data to correct a differential misclassification error
identified in a subsequent analysis of the study data. An association has also been reported for
exposure to bromodichloromethane (or a closely associated compound) and 1) stillbirth of fetuses
weighing more than 500 g and 2) increased risk of neural tube defects in women exposed to >20
|ig/L of bromodichloromethane prior to conception through the first month of pregnancy. An
association has been reported for total brominated trihalomethanes and reduced menstrual cycle
and follicular phase length in women of child-bearing age. Among the individual brominated
trihalomethanes, dibromochloromethane displayed the strongest association with altered
menstrual function. However, because chlorinated water contains many disinfection by-products,
it is not possible to directly conclude from these studies that bromodichloromethane or
bromodichloromethane are developmental toxicants in humans. Nevertheless, these studies raise
significant concern for possible human health effects. The methodology used to estimate
exposure to brominated trihalomethanes in tap water has been examined with the goal of refining
estimates of intake of these compounds in epidemiological studies.
Susceptible Populations
There is currently no clear evidence to suggest that children or the fetus are at greater risk
for adverse effects from exposure to bromoform or dibromochloromethane than are adults.
Where evidence of developmental toxicity has been observed in animals exposed to these
chemicals, it generally appears to occur only at doses that elicit signs of maternal toxicity.
Associations between concentration of bromodichloromethane (or a co-occurring chemical) and
spontaneous abortion or still birth have been observed in two epidemiological studies. Evidence
in rats indicates that exposure to bromodichloromethane causes full litter resorption in F344 rats
but not Sprague-Dawley rats, at doses many-fold higher than are likely to occur from human
ingestion of drinking water. Full litter absorption appears to result from a maternally-mediated
mode of action, rather than from a direct effect on the embryo. A mechanism of action for
bromodichloromethane-related pregnancy loss is suggested for the rat (reduced sensitivity of the
corpus luteum to luteinizing hormone), but is not without alternative explanation. At present,
there is insufficient information on the mode of action leading to full litter resorption in rats to
fully evaluate the relevance of this outcome to potential reproductive and/or developmental
toxicity in humans. There is presently no evidence to suggest that or the fetus are at increased
risk for brominated trihalomethane toxicity as a result of higher levels of metabolizing enzymes.
Subpopulations with either high levels of glutathione ^'-transferase or low baseline levels
of glutathione may potentially be more sensitive than the general population to brominated
trihalomethane-induced toxicity, but there are currently no epidemiological or animal data to
confirm this speculation. The functional significance of polymorphisms in cytochrome P450
isoforms that metabolize brominated trihalomethanes is also unknown. Populations that have
higher levels of CYP2E1 activity as a result of co-exposure to ethanol or other inducers or as a
result of altered health or physiological states may experience potentially higher risk.
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Mechanism of Toxicity
It is generally believed that the toxicity of the brominated trihalomethanes is related to
their metabolism. This conclusion is based largely on the observation that liver and kidney, the
chief target tissues for these compounds, are also the primary sites of their metabolism. In
addition, treatments which increase or decrease metabolism also tend to increase or decrease
trihalomethane-induced toxicity in parallel.
Metabolism of brominated trihalomethanes is believed to occur via oxidative and reductive
pathways. Limited structure-activity data for brominated trihalomethanes and the structurally-
related trihalomethane chloroform suggest that bromination may influence the proportion of
compound metabolized via the oxidative and reductive pathways, with brominated compounds
being more extensively metabolized by the reductive pathway. Additional evidence suggests that
a GSH-mediated pathway may play an important role in metabolism of brominated
trihalomethanes. These data raise the possibility that brominated trihalomethanes may induce
adverse effects (toxicity and carcinogenicity) via several different pathways.
The precise biochemical mechanisms which link brominated trihalomethane metabolism to
toxicity have not been characterized, but many researchers have proposed that toxicity results
from the production of reactive intermediates. Reactive intermediates may arise from either the
oxidative (dihalocarbonyls) or the reductive (free radicals) pathways of metabolism. Such
reactive intermediates are known to form covalent adducts with various cellular molecules, and
may impair the function of those molecules and cause cell injury. Free radical production may
also lead to cell injury by inducing lipid peroxidation in cellular membranes. Direct evidence
showing a relationship between the level of covalent binding intermediates generated by either
pathway and the extent of toxicity is not available for the brominated trihalomethanes.
Manipulation of cellular glutathione levels suggests that this compound may play a protective role
in brominated trihalomethane-induced toxicity.
Individual brominated trihalomethanes have been shown to induce tumors in laboratory
animals. The mode of action by which brominated trihalomethanes induce tumors in target tissues
has not been fully characterized. DNA adducts can be formed by interaction of reactive
metabolites (resulting from oxidative and reductive metabolism) with DNA. In addition,
preliminary evidence suggests that DNA adducts can be formed through conjugation with
glutathione and bioactivation of the resulting conjugates. In contrast to chloroform, the role of
cytotoxicity and associated regenerative cell proliferation in tumorigenicity of brominated
trihalomethanes is presently unclear. Comparison of dose-response data for liver toxicity
(including cell proliferation) and tumorigenicity in mice suggests that tumor formation occurs at
concentrations lower than those which stimulate cell proliferation. No evidence for increased cell
proliferation in kidney was obtained in studies using doses up to 246 mg/kg-day for
bromodichloromethane, 312 mg/kg-day for dibromochloromethane, or 379 mg/kg-day for
bromoform.
Interaction with agents which increase or decrease the activity of enzymes responsible for
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metabolism of brominated trihalomethanes may modify carcinogenicity/toxicity. Pretreatment
with inducers of CYP2E1 has been observed to increase the hepatotoxicity of
bromodichloromethane and dibromochloromethane in male rats. Pretreatment with m-xylene, an
inducer of the CYP2B1/CYP2B2 isoforms, increased the hepatotoxicity of
dibromochloromethane in male rats. Conversely, administration of the cytochrome P450 inhibitor
1-aminobenzotriazole prevented bromodichloromethane-induced hepatotoxicity in rats. Recent
findings indicating possible glutathione-mediated metabolism of brominated trihalomethanes
suggest that treatments or agents which alter glutathione-»Y-transferase activity could potentially
modify the toxicity of brominated trihalomethanes.
The severity of brominated trihalomethane toxicity is potentially affected by the vehicle of
administration. In a study of vehicle effects on the acute toxicity of bromodichloromethane, a
high dose (400 mg/kg) of the chemical was more hepato- and nephrotoxic when given in corn oil
compared to aqueous administration, but this difference was not evident at a lower dose (200
mg/kg).
Quantification of Noncarcinogenie Effects
Candidate health effects endpoints were analyzed by benchmark dose (BMD) modeling
using a benchmark response of 10% added risk. The BMDL10 was defined as the 95% lower
bond on the BMD estimate. For bromodichloromethane, a BMDL10 of 30 mg/kg-day identified
on the basis of full litter resorption in F344 rats was used to calculate a One-day Health Advisory
(HA) value of 1 mg/L. A BMDL10 of 18 mg/kg-day for single cell hepatic necrosis, identified in a
30-day drinking water study in rats, was used to calculate a Ten-day HA value of 0.6 mg/L. A
BMDL10 of 18 mg/kg-day for reduced maternal body weight gain on gestation days 6-9, identified
in a developmental study in rats, was used to calculate a Longer-term HA of 0.6 mg/L for a 10-kg
child. A Longer-term HA value of 2 mg/L was calculated for a 70-kg adult based on the same
endpoint. The calculations for the Reference Dose (RfD) of 0.003 mg/kg-day and Drinking
Water Exposure Level (DWEL) of 90 |ig/L employed a duration adjusted BMDL10 of 0.8 mg/kg-
day for fatty degeneration of the liver, identified in a 24 month dietary study in rats. Because
bromodichloromethane is classified as a probable human carcinogen, a Lifetime HA is not
recommended.
For dibromochloromethane, no suitable study was located for the calculation of a One-day
HA value. Use of the 10-day HA value as a conservative estimate is recommended. The Ten-day
HA value of 0.6 mg/L was calculated using a BMDL10 of 5.5 mg/kg-day for hepatic cell
vacuolization, identified in a 28-day feeding study in rats. A duration-adjusted BMDL10 value of
1.7 mg/kg-day for hepatic cell vacuolization, identified in a 13-week gavage study in rats, was
used to calculate Longer-term HA values of 0.2 and 0.6 mg/L for a 10-kg child and a 70-kg adult,
respectively. A duration-adjusted BMDL10 value of 1.6 mg/kg-day for fatty changes identified in
a 2 year gavage study in mice was used to calculate a RfD of 0.02 mg/kg-day and a DWEL of 700
|ig/L. The Lifetime HA for dibromochloromethane is 60 |ig/L. This value was calculated using
the default RSC value of 80% for exposure via ingestion of drinking water. Because this
compound is classified as a possible human carcinogen, the derivation of the Lifetime HA
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incorporated an uncertainty factor of 10.
For bromoform, an estimated dose of 54 mg/kg-day that caused sedation in children was
used to calculate a One-day HA value of 5 mg/L. A BMDL10 of 2.3 mg/kg-day for hepatic
vacuolization, identified in a one month dietary study in rats, was used to calculate a value of
0.2 mg/L for the Ten-day HA for the 10-kg child. This value was also recommended for use as
the Longer-term HA for a 10 kg child. A duration-adjusted BMDL10 value of 2.6 mg/kg-day for
hepatic vacuolization, identified in a 13 week gavage study in rats, was also used to calculate a
value of 0.9 mg/L for the Longer-term HA for the 70-kg adult. The BMDL10 value of 2.6 mg/kg-
day was also used to calculate an RfD of 0.03 mg/kg-day and a DWEL of 1000 |ig/L. Because
bromoform is classified as a probable human carcinogen, a Lifetime HA is not recommended.
Quantification of Carcinogenic Effects
Chronic oral exposure studies performed by the National Toxicology Program in rats and
mice provide adequate data to derive quantitative cancer risk estimates for the three brominated
trihalomethanes, although the chemicals were administered in a corn oil vehicle. For bromodi-
chloromethane, a unit risk of 1.0 x 10"6 (|ig/L)"' was derived, based on the incidence of renal
tumors in male mice. The oral slope factor and concentration for excess cancer risk of 10"6 were
3.5 x 10"2 (mg/kg-day)"1 and 1.0 |ig/L, respectively. For dibromochloromethane, a unit risk of 1.2
x 10"6 (|ig/L)"' was derived, based on liver tumors in female mice. The oral slope factor and
concentration for excess cancer risk of 10"6 were 4.3xlO"2 (mg/kg-day)"1 and 0.8 |ig/L,
respectively. For bromoform, a unit risk of 1.3 x 10"7 (|ig/L)"' was derived, based on tumors of the
large intestine in female rats. The oral slope factor and concentration for excess cancer risk of 10"
6 were 4.56><10"3 (mg/kg-day)"1 and 8 |ig/L, respectively. These values were calculated using an
animal-to-human scaling factor of body weight34 in accordance with proposed U.S. EPA guidance
(U.S. EPA, 1996; 1999).
In a previous assessment of the carcinogenicity of brominated trihalomethanes, the
Carcinogenic Risk Assessment Verification Endeavor (CRAVE) group of the U.S. EPA assigned
bromodichloromethane and bromoform to Group B2: probable human carcinogen. CRAVE
assigned dibromochloromethane to Group C: possible human carcinogen. Under the proposed
1999 U.S. EPA Guidelines for Cancer Assessment, bromodichloromethane and bromoform are
likely to be carcinogenic to humans by all routes of exposure. This descriptor is appropriate
when the available tumor data and other key data are adequate to demonstrate carcinogenic
potential to humans. This finding is based on the weight of experimental evidence in animal
models which shows carcinogenicity by modes of action that are relevant to humans.
Dibromochloromethane shows suggestive evidence of carcinogenicity, but not sufficient to assess
human carcinogenic potential. This descriptor is used when the evidence from human or animal
data is suggestive of carcinogenicity, which raises a concern for carcinogenic effects but is not
judged sufficient for a conclusion as to human carcinogenic potential. This finding is based on the
weight of experimental evidence in animal models which indicate limited or equivocal evidence of
carcinogenicity.
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IARC has recently re-evaluated the carcinogenic potential of the brominated
trihalomethanes. IARC classified bromodichloromethane as a Group 2B carcinogen: possibly
carcinogenic to humans. IARC classified dibromochloromethane and bromoform as Group 3: not
classifiable as to carcinogenicity in humans.
Table 1-1 summarizes the quantification of noncarcinogenic and carcinogenic effects for
brominated trihalomethanes.
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Table 1-1 Summary of Quantification of Toxicological Effects for Brominated
Trihalomethanes
Advisory
Value
Reference
Bromodichloromethane
One-day HA for 10-kg child
1 mg/L
Narotsky et al. (1997)
Ten-day HA for 10-kg child
0.6 mg/L
NTP (1998)
Longer-term HA for 10-kg child
0.6 mg/L
CCC (2000d)
Longer-term HA for 70-kg adult
2 mg/L
CCC (2000d)
RID
0.003 mg/kg-day
Aida et al. (1992b)
DWEL
100 ng/L
Aida et al. (1992b)
Lifetime HA
Not applicable
-
Oral Slope Factor c
3.5 x 10"2 (mg/kg-day)"1
NTP (1987)
Concentration for excess cancer risk (10"6)
1.0 ng/L
NTP (1987)
Unit Risk
lxlO"6 (mg/L)"1
NTP (1987)
Dibromochloromethane
One-day HA for 10-kg childb
0.6 mg/L
Aida et al. (1992a)
Ten-day HA for 10-kg child
0.6 mg/L
Aida et al. (1992a)
Longer-term HA for 10-kg child
0.2 mg/L
NTP (1985)
Longer-term HA for 70-kg adult
0.6 mg/L
NTP (1985)
RfD
0.02 mg/kg-day
NTP (1985)
DWEL
700 ng/L
NTP (1985)
Lifetime HA
60 (ig/L
NTP (1985)
Oral Slope Factor c
4.3 x 10"2 (mg/kg-day)-1
NTP (1985)
Concentration for Excess cancer risk (10"6)
0.8 ng/L
NTP (1985)
Unit Risk
1.2 x 10"6 (ng/L)"1
NTP (1985)
Bromoform
One-day HA for 10-kg child
5 mg/L
Burton-F arming (1901)
Ten-day HA for 10-kg child
0.2 mg/L
NTP (1989a)
Longer-term HA for 10-kg child a
0.2 mg/L
NTP (1989a)
Longer-term HA for 70-kg adult
0.9 mg/L
NTP (1989a)
RID
0.03 mg/kg-day
NTP (1989a)
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Table 1-1 (cont.)
Advisory
Value
Reference
DWEL
300 ng/L
NTP (1989a)
Lifetime HA
Not applicable
--
Oral Slope Factor c
4.6*10"3 (mg/kg-day)"1
NTP (1989a)
Concentration for Excess cancer risk (10"6)
8 ng/L
NTP (1989a)
Unit Risk
1.3 x 10"7 (ng/L)"1
NTP (1989a)
a The calculated value for the Longer-term HA was slightly higher than the values for the Ten-day HAs.
Therefore, use of the Ten-day HA for a 10-kg child is recommended as an estimate of the Longer-term HA for a
10-kg child.
b Use of the Ten-day HA recommended as an estimate of the One-day HA for a 10-kg child.
c The oral slope factor was calculated using the Linearized Multistage model (extra risk) and an animal-to-human
scaling factor of body weight3 4
Abbreviations: BW, Body weight; DWEL, Drinking water exposure limit; HA, Health advisory; LMS; Linearized
Multistage Model
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II. PHYSICAL AND CHEMICAL PROPERTIES
A. Properties and Uses
Bromodichloromethane (CHBrCl2), dibromochloromethane (CHBr2Cl) and bromoform
(CHBr3) are clear liquids with higher densities than the structurally-related compound chloroform.
They have limited solubility in water but are very soluble in organic solvents (Windholz, 1976).
Some important physical and chemical properties of these bromine-containing trihalomethanes are
summarized in Table II-1. Brominated trihalomethanes are sufficiently volatile to evaporate from
drinking water (Jolley et al., 1978).
Table II-l Physical and Chemical Properties of the Brominated Trihalomethanes
Property
Chemical
Bromodichloromethane
Dibromochloromethane
Bromoform
Structure
H
CI C CI
Br
H
CI C Br
Br
H
Br C Br
Br
Chemical Abstracts
Registry Number
75-27-4
124-48-1
75-25-2
Registry of Toxic Effects of
Chemical Substances
Number
PA 5310000
PA 6360000
PB 5600000
Synonyms
dichlorobromomethane
chlorodibromomethane
tribromomethane
Chemical Formula
CHBrCl2
CHBr2Cl
CHBr3
Molecular Weight
163.83
208.29
252.77
Boiling Point
90°C
116C
149 - 150°C
Melting Point
-57.1°C
--
6-7°C
Specific Gravity (20°)
1.980
2.38
2.887
Vapor Pressure
50mm (20°C)
15 mm (10°C)
5.6 mm (25°C)
Stability in Water
volatile
volatile
volatile
Water Solubility
3,032 mg/L
(30°C)
1,050 mg/L
(30°C)
3,190 mg/L (30°C)
Log Octanol: Water
Partition Coefficient (Kow)
2.09
2.23
2.37
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Source: USEPA (1992c; 1994b)
Brominated trihalomethanes also occur in drinking water as by-products of chlorination.
Bromide (Br"), a common constituent of natural waters, is oxidized by hypochlorous acid (HOCl3)
to form hypobromous acid (HOBr) in the following reaction:
Br + HOCI3 HOBr + CI"
Hypobromous acid reacts with naturally occurring organic substances in water (e.g.,
humic and fulvic acids) to yield the bromine-containing trihalomethanes bromoform,
dibromochloromethane and bromodichloromethane (in increasing order of formation rates) (Jolley
et al., 1978).
Trihalomethanes may also be produced by reaction with endogenous organic material in
the gut. Mink et al. (1983) treated adult male Sprague-Dawley rats with a single oral dose of
48 mg CI (as sodium chloroacetate) and 32 mg Br" (as potassium bromide). All three brominated
trihalomethanes were detected in the stomach contents of nonfasted rats following treatment
(Mink et al., 1983). Bromoform and dibromochloromethane were also detected in the plasma.
In the past, bromoform, bromodichloromethane and dibromochloromethane have been
used in pharmaceutical manufacturing and chemical synthesis, as ingredients in fire-resistant
chemicals and gauge fluids, and as solvents for waxes, greases, resins, and oils (U.S. EPA, 1975).
However, use patterns have changed over time. At present, the primary use of
bromodichloromethane is as a chemical intermediate for organic synthesis and as a laboratory
reagent (ATSDR 1989). Dibromochloromethane is reportedly used in laboratory quantities only
(AT SDR 1990). Use of bromoform is limited to performance of geological tests, use as a
laboratory reagent, and use in quality assurance programs in the electronics industry (ATSDR
1990).
B. Summary
Brominated trihalomethanes are volatile organic liquids that occur in drinking water as by-
products of disinfection with chlorine. The brominated trihalomethanes occurring in water are
bromoform, dibromochloromethane and bromodichloromethane. These compounds are formed in
water when hypochlorous acid oxidizes bromide ions to form hypobromous acid, which
subsequently reacts with organic material. In the past, individual brominated trihalomethanes
have been used for a variety of industrial purposes. Currently, these compounds are used as
laboratory reagents and, in the case of bromodichloromethane, as an intermediate in chemical
synthesis.
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III. TOXICOKINETICS
This section summarizes available information on the absorption, distribution, metabolism
and excretion of brominated trihalomethanes. Because the toxicokinetic properties of brominated
trihalomethanes appear to be generally similar, data in this section are presented for this class of
compounds as a group, rather than by individual chemical.
A. Absorption
Mink et al. (1986) compared the absorption of bromodichloromethane,
dibromochloromethane, and bromoform in male Sprague-Dawley rats and male B6C3F, mice.
The study animals received single oral doses of 14C-labeled compound in corn oil by gavage at
dose levels of 100 mg/kg (rats) or 150 mg/kg (mice). Total recovery of label in exhaled air, urine,
or tissues after 8 hours ranged from 62% to 93% (Table III-l), indicating that gastrointestinal
absorption was high for all three compounds. The level of radiolabeled carbon monoxide in
exhaled air was not quantified in this experiment. Carbon monoxide has since been recognized as
a product of brominated trihalomethane catabolism.
Mathews et al. (1990) administered 14C-bromodichloromethane by gavage in corn oil to
male Fischer 344 rats at doses of 1, 10, 32, or 100 mg/kg, and monitored the radiolabel in exhaled
air, urine, feces, and tissues. Absorption was extensive, with at least 86% of the dose recovered
as expired volatiles, C02, or CO. Only small amounts were recovered in urine (<5%) or in feces
(<3%) within 24 hours of administration, regardless of the size of the dose (Table III-2).
Table III-l Recovery of Label 8 Hours after Oral Administration of 14C-Labeled
Brominated Trihalomethanes to Male Sprague-Dawley Rats or Male B6C3Ft Mice
Chemical
Percent of Label
Species
Expired
co2
Expired
Parent
Urine
Organs
Total
Recovery
Bromodichloromethane
Rat
14.2
41.7
1.4
3.3
62.7
Mouse
81.2
7.2
2.2
3.2
92.7
Dibromochloromethane
Rat
18.2
48.1
1.1
1.4
70.3
Mouse
71.6
12.3
1.9
5.0
91.6
Bromoform
Rat
4.3
66.9
2.2
2.1
78.9
Mouse
39.7
5.7
4.6
12.2
62.2
Adapted from Mink et al. (1986) and U.S. EPA (1994b).
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Table III-2 Cumulative Excretion of Label after Oral Administration of 14C-Labeled
Bromodichloromethane to Male F344 Rats
Dose
Time
(hrs post-
treatment)
Percent of Dose
Expired
co2
Expired
CO
Expired
Volatiles
Urine
Feces
Total
Recovery
1 mg/kg
1
9.5±1.1
NRa
2.1il.5
NR
NR
11.6il.3
4
37.0±3.2
1.5i0.7
2.7il.8
NR
NR
41.1i2.8
8
62.9±2.2
2.7il.l
NR
NR
NR
68.il.7
16
76.4±3.2
NR
NR
NR
NR
81.8i2.9
24
77.5±3.3
3.3il.5
3.Oil.6
4.1i0.2
2.7il.5
90.7il.8
10 mg/kg
1
8.0±2.0
NR
2.0i0.8
NR
NR
10.0il.85
4
39.9±3.2
1.9i0.4
2.7il.l
NR
NR
44.5i3.0
8
66.0±4.0
3.4i0.9
NR
NR
NR
72.1i3.9
16
81.3±1.7
NR
NR
NR
NR
87.4il.5
24
82.1±1.8
4.3il.O
2.8il.l
4.3i0.2
0.7i0.2
94.2il.6
100
mg/kg
1
1.9±0.9
O.liO
1.5il.2
NR
NR
4.6il.8
4
5.5±1.8
0.3i0.1
4.2il.9
NR
NR
10.0i2.9
8
NR
NR
NR
0.6i0.4
NR
10.6i3.0
16
33.4±7.4
2.3i0.7
5.7i2.1
NR
NR
42.0i8.3
24
71.0±1.7
5.2i0.3
6.3i2.1
4.1i0.2
0.7i0.3
87.3il.6
Adapted from Mathews et al. (1990) and U.S. EPA (1994b).
aNot reported
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Lilly et al. (1998) examined the effects of vehicle on the absorption of orally administered
bromodichloromethane in an experiment designed to develop and validate a physiologically-based
pharmacokinetic model. Male F344 rats (3 animals/dose/vehicle/assay) were gavaged with 0, 50,
or 100 mg bromodichloromethane/kg in either corn oil or 10% Emulphor®, and
bromodichloromethane levels were monitored in blood and exhaled air. The dose levels
approximated doses previously utilized in two-year cancer bioassays of bromodichloromethane
(NTP, 1987). Concentrations of bromodichloromethane in blood and exhaled air peaked rapidly,
reaching maximal concentrations less than one hour after administration. The vehicle of
administration had significant effects on the blood and exhaled air concentrations. Delivery of
bromodichloromethane in 10% Emulphor® resulted in faster initial uptake, as inferred from higher
blood, tissue and breath chamber concentrations, when compared to corn oil (data presented
graphically). At 6 hours after administration, more than 90% and 100% of the administered dose
had been absorbed from the corn oil and Emulphor® vehicles, respectively.
B. Distribution
Data on the distribution of brominated trihalomethanes in exposed humans are limited.
Roth (1904) measured the bromoform content of tissues of a man who died from an accidental
oral overdose of bromoform and found levels in stomach and lung of 130 and 90 mg/kg wet
weight, respectively. Lower levels were reported in the intestine, liver, kidney, and brain.
Pellizzari et al. (1982) measured trihalomethanes in 42 samples of human milk taken from women
in urban areas. Dibromochloromethane was detected in one sample. Neither the level measured
nor the detection limit were reported for this study.
Data on the distribution of brominated trihalomethanes in animals are available from
studies in rats and mice. Mink et al. (1986) compared the distribution of bromodichloromethane,
dibromochloromethane, and bromoform in male Sprague-Dawley rats and male B6C3F, mice.
Single oral doses of 14C-labeled compound in corn oil were administered by gavage at dose levels
of 100 mg/kg (rats) or 150 mg/kg (mice). Tissue levels of radioactivity were measured 8 hours
after dose administration. The chemical form of the label measured in the tissues (e.g. parent or
metabolite, bound or free) was not determined. In the rat, the total organ content of label ranged
from 1.4%) to 3.6%> for the various compounds. The stomach, liver, and kidneys contained higher
levels than bladder, brain, lung, muscle, pancreas, and thymus. In mice, 4%> to 5%> of the
administered compound was recovered in the organs. However, an additional 10%> of the label
associated with bromoform was recovered in the blood of mice, yielding total organ levels of 12%>
to 14%). The authors attributed this elevated recovery of label to carboxyhemoglobin formation.
The levels of carboxyhemoglobin were not measured in this experiment.
Mathews et al. (1990) investigated the distribution of bromodichloromethane following
oral exposure in male Fischer 344 rats. Animals were given a single oral gavage dose of 1, 10,
32, or 100 mg/kg of 14C-bromodichloromethane dissolved in corn oil. Approximately 3%> to 4%>
of the administered dose was detected in tissues after 24 hours. The highest levels (1%> to 3%>)
were measured in liver. Repeated doses of 10 or 100 mg/kg-day for 10 days resulted in total
retention of only 0.9%> to 1.1% of the administered label, and had no effect on the tissue
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distribution of bromodichloro-methane.
The Chlorine Chemistry Council (CCC, 2000c) sponsored a study which analyzed the
levels of bromodichloromethane in parental tissues and fluids and Fx generation tissues as part of a
reproductive and developmental study in Sprague-Dawley rats (see Section V.E. 1 for a full study
description). Data from this study were summarized in Christian et al. (2001b).
Bromodichloromethane was administered in the drinking water at concentrations of 0, 50, 150,
450, or 1350 ppm. The estimated dosage on a mg/kg-day basis varied with the stage of the study
(see Table V-7). Plasma and other tissue samples were collected for analysis as described in
Table III-3. All samples were maintained frozen and shipped to the analytical lab (Lancaster
Laboratories, Lancaster, PA). Analysis of plasma collected from male and female rats prior to
mating and from female rats during gestation and lactation did not detect quantifiable levels of
bromodichloromethane (limit of detection 0.11 |ig). Bromodichloromethane was detected at a
concentration of 0.38 ju.g/g in the milk from one female in the 1350 ppm group.
Bromodichloromethane was not detected in placentas, amniotic fluid, or fetal tissue collected on
GD 21 or in plasma collected from postpartum day 29 weanling pups.
Table III-3 Over view of Tissue Collection for Analysis of Bromodichloromethane in
Sprague-Dawley Rat Tissues and Fluids (CCC, 2000c).
Generation
Sex
Physiological
state
Tissue
Day of
collection
No. of Samples;
Collection freq.
Comments
P
M, F
Pre-mating
Plasma
Day 1 of
exposure
3 rats/sex/group;
3 times/day
-
P
M, F
Pre-mating
Plasma
Day 14 of
exposure
3 rats/sex/group;
3 times/day
-
P
F
Pregnant
Plasma
GD 20
3 rats/group;
3 times/day
Rats con-
tinuously
exposed since
study day 1
P
F
Pregnant
Placenta
amniotic
fluid, and
fetuses
GD 21
3 litters/day
Tissues pooled
by litter
P
F
Lactating
Plasma
LD 15
3 rats/group
3 times/day
P
F
Lactating
Milk
LD 15
3 rats/group
1IU oxytocin
admin, by IV
approx. 5 min.
before milking
F,
M, F
Weaning
Plasma
LD 29
3 pups/sex;
3 litters;
3 times/day
"
Modified from CCC (2000c)
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Abbreviations: P, parental; M, male; F, female; GD, gestation day; LD, lactation day
The Chlorine Chemistry Council (CCC, 2000a) sponsored a study which analyzed the
levels of bromodichloromethane in parental tissues and fluids and Fx generation tissues as part of a
reproductive and developmental study in New Zealand White rabbits (see Section V.E. 1 for a full
study description). Data from this study were summarized in Christian et al. (2001b).
Bromodichloromethane was administered to groups of rabbits (4/concentration) in the drinking
water at concentrations of 0, 50, 150, 450, or 1350 ppm. The estimated doses at these
concentrations were 0, 4.9, 13.9, 32.3, or 76.3 mg/kg-day, respectively. Blood samples were
collected on GD 7 and 28. Amniotic fluid, and placenta samples were collected on GD 29 after
collection of a third blood sample, and amniotic fluid and placental tissue were pooled by litter.
Blood samples were collected from three randomly selected fetuses per litter. Bromodichloro-
methane was detected at concentrations of 0.15 and 0.17 |ig/g (limit of detection 0.11 jug/g) in
placentas from two litters in the 1350 ppm exposure group. Bromodichloromethane was detected
in one fetus from the 1350 ppm group "...at a level below the limit of detection".
Bromodichloromethane was not detected in placentas from does exposed to concentrations up to
450 ppm, in amniotic fluid from does exposed to concentrations up to 1350 ppm, or in the
remaining fetuses of does exposed to concentrations as high as 1350 ppm.
C. Metabolism
1. Overview
The toxicity of the brominated trihalomethanes is mediated by cytochrome P450-mediated
bioactivation to reactive metabolites. The pathways for brominated trihalomethane metabolism
were initially inferred from studies of the structurally-related trihalomethane chloroform (U.S.
EPA, 1994b). Additional details of brominated trihalomethane metabolism have subsequently
been elucidated in a number of laboratories using both in vitro and in vivo approaches and are
described below. Figure III-l presents a general metabolic scheme for chloroform and the
brominated trihalomethanes.
The metabolism of brominated trihalomethanes occurs via at least two pathways (U.S.
EPA 1994b). The oxidative pathway requires NADPH and oxygen, whereas the reductive
pathway can utilize NADPH or NADH and is inhibited by oxygen. Both reactions are believed to
be mediated by cytochrome P450 isoforms. In the presence of oxygen (oxidative metabolism),
the reaction product is trihalomethanol (CX3OH), which spontaneously decomposes to yield a
reactive dihalocarbonyl (CX20) such as phosgene (CC120). Dihalocarbonyls may undergo a
variety of reactions, such as adduct formation with various cellular nucleophiles, hydrolysis to
yield carbon dioxide, or glutathione-dependent reduction to yield carbon monoxide. When
oxygen levels are low (reductive metabolism), the metabolic reaction products appear to be free
radical species such as the dihalomethyl radical (*CHX2). These radicals are highly reactive and
may also form covalent adducts with a variety of cellular molecules. Evidence supporting this
metabolic scheme and information on species differences in the rate and extent of trihalomethane
metabolism are presented below. Additional data derived from studies of chloroform
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Figure III-l Proposed Metabolic Pathways for Brominated Trihalomethanes
RH
Reductive Pathway
< =»
P-450
•CHX, ^ CHX3
Oxidative Pathway
V2
X"
1 >
P-450
~
T
NADPH 2NADPH
or
NADH
O, H,0
- cx3oh ——CX20
HX
2GSH
i.
rcx2oh
-~ CO+GSSG
2HX
H,0
¦CO,
2HX
Cysteine
-~OTZ
2HX
X = halogen atom (chlorine or bromine); R = cellular nucleophile (protein, nucleic acid);
GSH = reduced glutathione; GSSG = oxidized glutathione;
OTZ = oxothiazolidine carboxylic acid; P-450 = cytochrome P-450
Adapted from Stevens and Anders (1981); Tomasi et al. (1985)
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are described in U.S. EPA (1994b).
The metabolism of trihalomethanes (including chloroform and the brominated
trihalomethanes) has been most intensively studied using chloroform as a substrate. These studies
indicate that many factors influence metabolism, including strain, species, chloroform
concentration, and possibly gender. A comprehensive review of chloroform studies is beyond the
scope of this document. However, because the P450-mediated metabolism of the brominated
trihalomethanes is expected to be similar to that of chloroform, descriptions of a few
representative studies of chloroform metabolism are provided to provide additional background
information on the metabolism of trihalomethanes.
A key question in hazard characterization is the identity of the P450 isoforms responsible
for bioactivation of brominated trihalomethanes to reactive metabolites. This is because
individuals or subpopulations with elevated levels of these enzymes may be at greater risk for
adverse effects. The identities of the cytochrome P450 isoforms responsible for trihalomethane
metabolism have been investigated most intensively in studies of chloroform (studies of
brominated trihalomethanes are described in sections 2 and 3 below).
Studies by Nakajima et al. (1995) and Testai et al. (1996) indicate that chloroform
concentration plays a critical role in determining the role of different isoforms and the associated
effects of metabolic inducers. Nakajima et al. (1995) pretreated male Wistar rats with three
inducers of specific P450 isoforms and subsequently administered a single dose of chloroform by
gavage in corn oil. The inducers used were phenobarbital (CYP2B1/2), n-hexane (CYP2E1), and
2-hexanone (CYP2B1/2 and CYP2E1). Liver damage (as determined by serum enzyme activity
and histopathology) was greatest at the mid-dose in the hexane-treated animals. In contrast, rats
pretreated with phenobarbital or 2-hexanone showed a dose-related increase of liver damage at all
dose levels. The pattern of damage was consistent in each case with the tissue distribution
patterns of the induced cytochrome P450 isoform(s). The study authors concluded on the basis of
these results that CYP2E1 catalyzes chloroform metabolism at low doses and that CYP2B1/2
catalyzes chloroform metabolism at higher doses.
While experimental evidence indicates that CYP2E1 and CYPB1/2 catalyze the oxidative
pathway, the identities of the cytochrome P450 isoforms which catalyze the reductive pathway
have not been established. In general, CYP2E1 protein can catalyze reductive as well as oxidative
reactions (Lieber, 1997) and this isoform has been implicated in the production of trichloromethyl
radicals from carbon tetrachloride (see Lieber et al. 1997). However, evidence for a dual role of
either CYP2E1 or CYP2B1/2 in catalyzing the oxidative and reductive pathways for
trihalomethane metabolism has been contradictory, perhaps as a result of the different
concentrations of chloroform used in different experiments (summarized in Testai et al. 1996).
To address the issue of concentration, Testai et al. (1996) studied the role of different isoforms in
chloroform using microsomes prepared from Sprague-Dawley rats pretreated with a variety of
cytochrome P450 inducers. The microsomes were incubated under varying conditions of
chloroform concentration, oxygenation, and presence of isoform-specific inhibitors or antibodies.
Under the conditions utilized in this series of experiments, the authors concluded that the
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cytochrome P450 isoforms involved in oxidative metabolism of brominated trihalomethanes do
not participate in the reductive pathway.
Recent evidence from studies in Salmonella typhimurium strains engineered to express the
rat glutathione ^-transferase theta 1-1 (GSTT1-1) gene suggests that bioactivation of brominated
trihalomethanes to mutagenic species is also mediated by one or more glutathione »S'-transferase-
mediated conjugation pathways (Pegram et al., 1997; DeMarini et al., 1997). Details of these
studies are presented in Section V.E. Proposed routes for glutathione conjugation and
bioactivation are illustrated in Figure V-2, located in Section V.E.
2. In Vitro Studies
Ahmed et al. (1977) investigated the in vitro oxidative (aerobic) metabolism of brominated
trihalomethanes to carbon monoxide by the rat liver microsomal fraction. Metabolism of
bromoform resulted in the highest level of carbon monoxide formation, followed by
dibromochloromethane and bromodichloromethane in decreasing order. Glutathione, NADPH
and oxygen were required for maximal carbon monoxide production. This activity was inducible
by phenobarbital or 3-methylcholanthrene pretreatment (agents which are known to increase
cytochrome P-450 activity) and was inhibited by the cytochrome P-450 inhibitor SKF 525-A.
Similar results were reported by Stevens and Anders (1979). In addition, Stevens and Anders
(1979) reported the formation of 2-oxothiazolidine-4-carboxylic acid (OTZ) when bromoform
was incubated in the presence of cysteine. Dihalocarbonyls react with cysteine to form OTZ.
Thus, detection of OTZ provides evidence that a dihalocarbonyl intermediate was formed during
bromoform metabolism.
Wolf et al. (1977) studied the in vitro metabolism of bromoform and chloroform to carbon
monoxide under anaerobic conditions using liver preparations from phenobarbital-induced rats.
Bromoform metabolism resulted in much greater levels of carbon monoxide production than did
the metabolism of chloroform. Gao and Pegram (1992) reported that binding of reactive
intermediates to rat hepatic microsomal lipid and protein under reductive (anaerobic) conditions
was more than twice as high for bromodichloromethane as for chloroform. These data suggest
that reductive metabolism may be a more important pathway for metabolism of brominated
trihalomethanes than for chloroform.
Tomasi et al. (1985) studied the anaerobic activation of bromoform,
bromodichloromethane, and chloroform to free radical intermediates in vitro using rat hepatocytes
isolated from phenobarbital-induced male Wistar rats. The production of a free radical
intermediate was measured by electron spin resonance (ESR) spectroscopy using the spin trap
compound phenyl-t-butylnitrone. The intensity of the ESR signal was greatest for bromoform,
followed by bromodichloromethane and then chloroform. The largest ESR signal was detected
when hepatocytes were incubated under anaerobic conditions. Incubation in the presence of air
resulted in a reduction of the signal, as did addition of cytochrome P-450 inhibitors such as SKF-
525A, metyrapone, and carbon monoxide. These data were interpreted to indicate that free-
radical formation depended on reductive metabolism of the trihalomethanes mediated by the
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cytochrome P450 system. Comparison of the ESR spectra for chloroform, deuterated
chloroform, and bromodichloromethane indicated that the free radical intermediate produced by
chloroform metabolism was *CHC12. The authors speculated that the brominated trihalomethanes
are also metabolized by transfer of an electron directly from the cytochrome to the halocompound
with the successive formation of the dihalomethyl radical (*CHX2) and a halide ion (X").
3. In Vivo Studies
Mink et al. (1986) compared the metabolic products of bromodichloromethane,
dibromochloromethane, and bromoform in male Sprague-Dawley rats and male B6C3F, mice
(strain not reported). Animals were given a single oral dose of 14C-labeled compound by gavage
in corn oil at dose levels of 100 mg/kg for rats and 150 mg/kg for mice. Levels of 14C were
measured in exhaled carbon dioxide recovered within 8 hours after dose administration. Expired
carbon dioxide accounted for 4.3% to 18.2% of the administered label in rats (Table III-l),
suggesting that the parent compound had undergone limited metabolism and oxidation. In mice,
the fraction of label excreted as carbon dioxide was higher, ranging from 40% to 81%. These
data indicate that oxidative metabolism of brominated trihalomethanes to carbon dioxide was
more rapid and extensive (by a factor of four- to ninefold) in mice than in rats. As previously
noted in Section III. A, production of carbon monoxide, a known metabolite of brominated
trihalomethanes, was not measured in this study.
Anders et al. (1978) investigated the formation of carbon monoxide from brominated
trihalomethanes in corn oil administered to Sprague-Dawley rats at doses of 1 mmol/kg (119 to
252 mg/kg) by intraperitoneal injection. Bromoform produced the highest levels of blood carbon
monoxide, followed by dibromochloromethane and bromodichloromethane in decreasing order.
A dose-response relationship was noted for bromoform following administration of 252, 506, or
1,012 mg/kg. Carbon monoxide production was inducible by pretreatment with phenobarbital,
but pretreatment with 3-methylcholanthrene had no effect. Carbon monoxide production was
significantly inhibited by SKF-525-A. Administration of 3H-bromoform resulted in decreased
carbon monoxide formation when compared to bromodichloromethane and
dibromochloromethane, indicating that the carbon-hydrogen bond breakage may be the rate-
limiting step under aerobic conditions. Similar results were later reported by Stevens and Anders
(1981).
Tomasi et al. (1985) studied the in vivo metabolism of chloroform, bromodichloro-
methane, and bromoform to free radical intermediates in rats. Starved, phenobarbital-induced
male Wistar rats (number not stated) were given intraperitoneal injections of 1,100 mg/kg
chloroform, 820 mg/kg bromodichloromethane, or 1,260 mg/kg bromoform dissolved in olive oil.
The animals were sacrificed and the livers were homogenized. The production of a free radical
intermediate by the livers was determined by ESR spectroscopy. The authors reported detection
of free radicals in the livers of all treated rats. The intensity of the ESR signal followed a ranking
similar to that observed in in vitro experiments (bromoform > bromodichloromethane >
chloroform), confirming that the reductive formation of free radicals is greater for brominated
trihalomethanes than for chloroform.
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Mathews et al. (1990) studied the metabolism of 14C-bromodichloromethane in male
Fischer 344 rats. Animals (n = 4) were given a single oral dose of 1, 10, 32, or 100 mg/kg of
bromodichloromethane dissolved in corn oil. Levels of labeled carbon dioxide and carbon
monoxide in exhaled air were measured for 24 hours. Approximately 70% to 80% of the dose
was metabolized and exhaled as 14C02 and 3% to 5% of the dose as 14CO. However, 14C02
production was slower following a single dose of 100 mg/kg than after the administration of a
single dose of 32 mg/kg or lower, suggesting saturation of metabolism. Repeated doses of
100 mg/kg-day for 10 days resulted in an increased rate of 14C02 production, compared with the
initial rate. The authors concluded on the basis of these data that bromodichloromethane may
induce its own metabolism.
Thornton-Manning et al. (1994) evaluated the effect of bromodichloromethane exposure
on cytochrome P450 isozyme activity in female F344 rats (6/dose). Gavage doses of 75 to 300
mg/kg-day were administered in a solution of 10% Emulphor® (an emulsifier) for five consecutive
days. Treatment resulted in decreased activity of the CYP1A and CYP2B isozymes. In contrast,
there was no effect on CYP2E1 activity.
Pankow et al. (1997) investigated the metabolism of dibromochloromethane in male
Wistar rats following single and repeated gavage doses. Rats receiving a single-dose (6 animals
per group) were treated with 0 (vehicle only), 0.4, 0.8, 1.6 or 3.1 mmol dibromochloro-
methane/kg dissolved in olive oil. Rats receiving multiple doses were gavaged with 0 (vehicle
only) or 0.8 mmol dibromochloromethane/kg once a day for 7 days. The blood or plasma
concentrations of parent compound, bromide, and carbon monoxide (as carboxyhemoglobin,
COHb) were measured following dibromochloromethane administration. The level of oxidized
glutathione (GSSG) in the liver of treated animals was also assayed. Oral administration of
dibromochloromethane resulted in a significant elevation of plasma bromide levels at all doses
tested. Bromide did not return to baseline levels even after 10 days. Repeated administration of
0.8 mmol dibromochloromethane/kg resulted in significantly higher plasma levels of bromide than
were measured following a single dose of 0.8 mmol/kg. COHb was also elevated in a dose-
dependent manner following either single or repeated administration of dibromochloromethane,
but returned to baseline levels within 24 hours after treatment. GSSG levels were significantly
increased at 12- and 24-hour time points following a single 0.8 mmol/kg dose (no other doses
were examined). Levels of GSSG returned to baseline levels by 48 hours after treatment.
Pankow et al. (1997) conducted additional experiments to determine whether reduced
glutathione (GSH) is a requirement for in vivo dibromochloromethane metabolism and to identify
P450 isozymes involved in the metabolism of dibromochloromethane. Pretreatment of rats with
buthionine sulfoximine (an agent which depletes GSH) reduced GSH concentrations as
anticipated and decreased the rate of bromide and COHb production. In contrast, pretreatment
with butylated hydroxyanisole (which increases GSH levels) increased the rate of bromide and
COHb production. These results suggest that GSH plays a role in dibromochloromethane
metabolism. Further studies were conducted to determine which cytochrome P450 isoform(s)
participate in the in vivo metabolism of dibromochloromethane. Simultaneous exposure to 0.8
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mmol dibromochloromethane/kg and diethyldithiocarbamate (a potent inhibitor of P450 isoform
CYP2E1) partially inhibited the production of bromide and COHb. In contrast, pretreatment with
isoniazid (an potent inducer of CYP2E1) increased formation of bromide and COHb. These
experiments indicate that CYP2E1 is at least partially responsible for dibromochloromethane
metabolism. Pretreatment with phenobarbital, an inducer of cytochrome P450 isoforms CYP2B1
and 2B2, increased the concentration of bromide in plasma, suggesting that CYP2B1 and 2B2
may also participate in the catabolism of dibromochloromethane. Pretreatment with m-xylene,
which induces both CYP2E1 and CYP2B1/2, resulted in higher bromide levels than inducers of
CYP2E1 (isoniazid) or CYP2B1/2 (phenobarbital) administered individually. Pankow et al.
(1997) concluded on the basis of these multiple experiments that 1) bromide and carbon
monoxide are metabolites of dibromochloromethane; 2) dibromochloromethane is metabolized via
the oxidative pathway; 3) the oxidative metabolism of dibromochloromethane is catalyzed by
CYP2E1 and CYP2B1/2; and 4) dibromochloromethane plays a role in the induction of CYP2E1.
Allis et al. (2001) investigated the effect of inhalation exposure to bromodichloromethane
on the activity and protein levels of CYP1A2, CYP2B1, and CYP2E1 in female F344 rats
(6/dose). In addition, the effect of inhalation exposure on the activity level of CYP1A1 was
assayed. Serum bromide ion concentration, an indicator of the total metabolism of
bromodichloromethane, was measured in samples drawn from a separate set of animals
(4/concentration) exposed to the same concentrations. The test animals were exposed for 4 hours
to measured bromodichloromethane concentrations of 0, 106, 217, 419, 812, 1620, and 3240
ppm. The microsomal isozyme activities assayed were: p-nitrophenol hydrolase (PNP), an
indicator of CYP2E1 activity; pentoxyresorufin-O-dealkylase (PROD), an indicator of CYP2B1/2
activity; ethoxsyresorufin-O- dealkylase (EROD), an indicator of CYP1 Al activity; and
methoxyresorufin-O-dealkylase (MROD), an indicator of CYP1A2 activity. The pattern of results
for isozyme activity obtained in this inhalation study was similar to that reported for male F344
rats treated with bromodichloromethane by gavage. CYP2E1 activity as measured by PNP
activity was not significantly affected by treatment. MROD, EROD, and PROD activities showed
modest increases at low exposure concentrations. The increases were statistically significant for
EROD and MROD at the 106 ppm exposure concentration. Decreases were observed at higher
exposure concentrations relative to controls. These decreases were statistically significant for
PROD at 3240 ppm and for EROD and MROD at concentrations of 800 ppm and greater. The
results for isozyme protein levels, as measured by Western blots, were generally consistent with
the results for isozyme activity. The study authors speculated that the most dramatic reductions
in isozyme activity (PROD and MROD) were a result of suicide inhibition. In addition, they
concluded that analyses of the EROD and MROD activity and protein level patterns indicates that
CYP1A2 is involved in the metabolism of bromodichloromethane. The study authors noted that it
is not typical for this isozyme to metabolize the small molecules (such as chloroform) that are the
usual substrates for CYP2E1, but observed that the presence of the large bromide ion may make
bromodichloromethane a suitable substrate for CYP1A2. Blood bromide concentration reached a
maximum at 200 ppm, indicating that metabolism was saturated a concentrations equal to or
greater than 200 ppm.
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Allis et al. (2002) reported additional evidence for metabolism of bromodichloromethane
by CYP1A2. Induction of CYP1A2 without parallel induction of CYP2E1 or CYP2B1/2
(accomplished by pretreatment with 2,3,7,8-tetrachlorodibenzodioxin, TCDD), increased
hepatotoxicity in male F344 rats administered a gavage dose of 400 mg/kg
bromodichloromethane. Hepatotoxicity was assessed by measurement of alanine
aminotransferase (ALT) and sorbitol dehydrogenase (SDH) activity. Pretreatment with TCDD
increased serum bromide levels (a measure of total bromodichloromethane metabolism) in rats
treated with 200 or 400 mg/kg when compared to uninduced controls. The apparent
inconsistency between lack of hepatotoxicity and increased total metabolism at 200 mg/kg was
explained by effective detoxification at this dose, presumably by glutathione. Selective inhibition
of CYP1A2 activity, by administration of isosafrole to TCDD-induced animals prior to treatment
with 400 mg/kg bromodichloromethane significantly reduced the hepatotoxic response and serum
bromide concentrations.
Allis and colleagues (Allis et al., 2002; Allis and Zhao, 2002; Zhao and Allis, 2002)
assessed the ability of various rat and human CYP isoenzymes to metabolize
bromodichloromethane and determined kinetic parameters for those showing measurable
metabolic activity. Allis and Zhao (2002) tested five rat and six human CYP isoenzymes in vitro
for metabolism of bromodichloromethane using recombinant systems expressing single isozyme
activities. The tested recombinant isoenzymes were rat CYP2E1, CYP2B1/2, CYP1A2,
CYP2C11, AND CYP3A1 and human CYP2E1, CYP1A2, CYP2A6, CYP2B6, CYP2D6 and
CYP3 A4. The results of this study indicate that the principal metabolizing enzymes in rat are
those identified previously, namely CYP2E1, CYP2B1/2, CYP1A2. Results for CYP3A1 suggest
that it may have weak metabolic activity, but the level of activity was not sufficient for a
quantitative assessment. CYP2C11 was not active. Human CYP2E1, CYP1A2, and CYP3A4
showed substantial metabolic activity toward bromodichloromethane. Human CYP2A6 showed
lower, but measurable, levels of activity. CYP2B6 and CYP2D6 were not active. Based on these
assays, only CYP2E1 and CYP1A2 metabolize bromodichloromethane in both species. CYP2E1
is the high affinity enzyme in both rats and humans, with Km values approximately 27-fold lower
than those for the isoenzymes with the next lowest value (CYP2B1 in rats, CYP1A2 in humans).
The metabolic parameters Km and &cat for rat and human CYP2E1 were similar. In contrast, the
metabolic parameters for CYP1A2 were not similar across species. The study authors concluded
that the results of this study appear consistent with observations in vivo for the rat (Allis et al.,
2002) and with predictions of the existing PBPK model for bromodichloromethane in the rat
(Lilly et al., 1998).
Zhao and Allis (2002) determined kinetic constants for metabolism of
bromodichloromethane by CYP2E1, CYP1A2, and CYP3A4 in human liver microsomes.
Constants for individual isoenzymes were determined by addition of enzyme-specific inhibitory
antibodies for two isoenzymes to the microsomal preparations while measuring the activity of the
third. Measurements were performed in microsomes obtained from four donors. CYP2E1 was
found to have the lowest Km(2.9 |iM) and the highest catalytic activity. The Km values for
CYP1A2 and CYP3A4 were approximately 20-fold higher (60 |iM) and the catalytic activity was
lower. Eleven additional human microsomal preparations were characterized for activity of 10
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CYP isoenzymes. The initial rate of metabolism in each preparation measured at 9.7 |iM
bromodichloromethane was compared to the activity of individual isoenzymes. Statistical analysis
showed a significant correlation only with CYP2E1 activity at the tested concentration. The
results of this suggest that CYP2E1 would dominate metabolism at the low level of exposure
expected from ingestion of drinking water.
, study However, the study authors have noted that humans are highly variable in the
induction of CYP isoenzymes and that the contributions of the three isoenzymes to metabolism of
bromodichloromethane in individuals may not be entirely predictable.
D. Excretion
Mink et al. (1986) compared the excretion of bromodichloromethane,
dibromochloromethane, and bromoform in male Sprague-Dawley rats and male B6C3F, mice.
Animals were given single oral doses of 14C-labeled compound in corn oil by gavage at dose levels
of 100 mg/kg and 150 mg/kg for rats and mice, respectively. The lung was the principal route of
excretion in both species, accounting for 45% to 88% of the administered label, either as carbon
dioxide or as parent compound. Small amounts of label (1.1% to 4.9%) were recovered in urine,
but the chemical identity of labeled compounds was not investigated.
Mathews et al. (1990) exposed Fischer 344 rats to either a single oral dose of 1, 10, 32, or
100 mg/kg, or 10-day repeated doses of 10 or 100 mg/kg-day bromodichloromethane dissolved in
corn oil. Approximately 70% to 80% of the administered dose was excreted in exhaled air as 14C-
carbon dioxide, with 3% to 5% as 14C-carbon monoxide. In general, less than 5% of the dose was
excreted in the urine or feces.
E. Bioaccumulation and Retention
No data were located regarding the bioaccumulation or retention of brominated
trihalomethanes following repeated exposures. However, based on the rapid excretion and
metabolism of the brominated trihalomethanes and the low levels of the structurally-related
compound chloroform detected in human post-mortem tissue samples, marked accumulation and
retention of these compounds are not anticipated.
F. Summary
No data on absorption of brominated trihalomethanes were available for humans.
Measurements in mice and rats indicate that gastrointestinal absorption of brominated
trihalomethanes is rapid (peak levels attained less than an hour after administration of a gavage
dose) and extensive (63% to 93%). Most studies of brominated trihalomethane absorption have
used oil-based vehicles. A study in rats found that the initial absorption rate of
bromodichloromethane was higher when the compound was administered in an aqueous vehicle
when compared to administration in a corn oil vehicle.
Data for distribution of brominated trihalomethanes in human organs and tissues are
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limited. Bromoform was found primarily in the stomach and lungs of a human overdose victim,
with lower levels detected in intestine, liver, kidney and brain. Dibromochloromethane was found
in 1 of 42 samples of human breast milk collected from women living in urban areas.
Radiolabeled brominated trihalomethanes were detected in a variety of tissues following oral
dosing in rats and mice. Approximately 1 to 4% of the administered dose was recovered in body
tissues when analysis was conducted 8 or 24 hours post-treatment. The highest concentrations
were detected in stomach, liver, blood, and kidneys when assayed 8 hours after administration of
the compounds. Analyses of placentas, amniotic fluid and fetuses from female rats and rabbits
administered bromodichloromethane in drinking water indicate that this compound does not
accumulate in these tissues or fluids. There are no data which are suggestive of strain specific
differences in metabolism.
Brominated trihalomethanes are extensively metabolized by animals. Metabolism of
brominated trihalomethanes occurs via two pathways. One pathway predominates in the presence
of oxygen (the oxidative pathway) and the other predominates under conditions of low oxygen
tension (the reductive pathway). In the presence of oxygen, the initial reaction product is
trihalomethanol (CX3OH), which spontaneously decomposes to yield the corresponding
dihalocarbonyl (CX20). The dihalocarbonyl species are quite reactive and may form adducts with
cellular molecules. When intracellular oxygen levels are low, the trihalomethane is metabolized
via the reductive pathway, resulting in a highly reactive dihalomethyl radical (*CHX2), which may
also form covalent adducts with cellular molecules. The metabolism of brominated
trihalomethanes and chloroform appear to occur via the same pathways, although in vitro and in
vivo data suggest that metabolism via the reductive pathway occurs more readily for brominated
trihalomethanes. Both oxidative metabolism and reductive metabolism of trihalomethanes appear
to be mediated by cytochrome P450 isoforms. The identity of cytochrome P450 isoforms that
metabolize brominated trihalomethanes has been investigated in several studies which used
bromodichloromethane as a substrate. The available data suggest that the cytochrome P450
isoforms CYP2E1, CYP2B1/2, and CYP1A2 metabolize bromodichloromethane in rats. The
human isoforms CYP2E1, CYP1A2, and CYP3A4 show substantial activity toward
bromodichloromethane in vitro and low but measurable levels of CYP2A6 activity have also been
detected. Based on the available data, CYP2E1 and CYP1A2 are the only isoforms active in both
rats and humans. CYP2E1 shows the highest affinity for bromodichloromethane in both species
and the metabolic parameters Km and kcal are similar for rat and human CYP2E1. In contrast, the
metabolic parameters for CYP1A2 differ in rats and humans. The pattern of results for isozyme
activity obtained from an inhalation study of bromodichloromethane was similar to the pattern
reported for male F344 rats treated with bromodichloromethane by gavage.
The lung is the principle route of excretion in rats and mice. Studies with 14C-labeled
compounds indicate that up to 88% of the administered dose can be found in exhaled air as
carbon dioxide, carbon monoxide, and parent compound. Excretion in the urine generally appears
to be 5% or less of the administered oral dose. Data from one study suggests that fecal excretion
is less than 3% of the administered dose.
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IV. HUMAN EXPOSURE
A. Occurrence in Drinking Water
The occurrence of brominated trihalomethanes in U.S. drinking water has been determined
in both national-scale and localized studies. The occurrence of bromodichloromethane and
bromoform has been described in eleven national surveys. Dibromochloromethane occurrence has
been described in twelve national surveys. Nine localized studies on the occurrence of brominated
trihalomethanes are also described below.
It is important to note that a variety of sampling and preservation techniques are used for
collection of occurrence data on brominated trihalomethanes. The addition of chlorine to raw
water as a disinfectant at water treatment plants results in the formation of hypochlorous acid in
the processed water. The acid in turn reacts with organic materials to produce chloroform and
also oxidizes available bromide ions to form hypobromous acid. Hypobromous acid reacts with
organic materials in the processed water to form the brominated trihalomethanes. Because these
chemical reactions occur over periods of days in treated waters, the method used to sample
drinking waters can affect the measured concentrations of trihalomethanes in the water.
Therefore, appropriate sampling and preservation methods must be selected to ensure that the
analytical data are representative of the desired endpoint. For example, if an investigator wants to
know the concentration of trihalomethanes in the water at the time of sampling, a reducing agent
is added to the sample containers to "quench" or prevent further formation of trihalomethanes. If
an investigator wants to know the maximum amount of trihalomethanes that could occur, no
quenching is used and the reactions are allowed to run to completion at room temperature. If a
concentration similar to that at a household tap is desired (i.e., after the water spends several days
in the distribution system, the samples generally are not quenched but are refrigerated to slow the
reactions (Wallace, 1997). Information on sample handling has been included in the discussion of
individual studies when available in the materials reviewed for this document.
Spatial and temporal variability exist in the occurrence data reported for brominated
trihalomethanes. Multiple factors contribute to this variability. With respect to spatial variability,
the geographical distribution of bromide ion in soil is not uniform (Shacklette and Boerngen,
1984). Brominated byproducts may predominate or comprise a substantial proportion of the
disinfection byproduct profile in regions with high soil concentrations. Brominated
trihalomethanes may continue to form within water distribution systems due to the action of free
residual chlorine on remaining humic precursors, resulting in substantial intra-system spatial
variability (Chen and Weisel, 1999). Temporal variability may result from seasonal variation in
the concentration of brominated trihalomethanes as a result of seasonal fluctuations in precursor
material (Brett et al., 1979). Short term variability may be introduced by changes in the demand
cycle to individual homes or neighborhoods.
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1. National Surveys
The National Organics Reconnaissance Survey (NORS), conducted by U.S. EPA,
collected drinking water samples from 80 cities nationwide in 1975. The survey sampled for
several organics, including brominated trihalomethanes, at the water treatment facilities. The
sampling method employed was refrigeration without quenching; therefore, brominated
trihalomethane concentrations may have increased following collection. Bromodichloromethane
was found in 98% of the systems sampled. The median concentration was 8 |ig/L (ppb), and the
maximum level was 116 |ig/L (ppb). Dibromochloromethane was found in 90% of the systems
sampled at a median concentration of 2 |ig/L (ppb). The detection limit for
dibromochloromethane and bromodichloromethane was 0.1 |ig/L (ppb). The median
concentration for bromoform was below the detection limit of approximately 5 |ig/L (ppb)
(Symons et al,. 1975). NORS was performed prior to the promulgation of the total
trihalomethane regulation; therefore, these results may be higher than current levels.
The National Organics Monitoring Survey (NOMS) was conducted by the EPA from
March 1976 to January 1977 (Wallace, 1997). In NOMS, 113 community water supplies were
sampled. Surface water was the major source for 92 of the systems, and ground water was the
major source for the remaining 21 systems. The NOMS used three sample storage methods.
During Phase 1, all samples were refrigerated. In Phase 2, the samples were allowed to stand at
20 to 25 °C for 2 to 3 weeks to maximize trihalomethane formation. Phase 3 had two parts. The
samples identified as 3T were allowed to stand an additional 2 to 3 weeks. The samples identified
as 3Q were quenched by addition of sodium thiosulfate. As expected, the highest trihalomethane
values occurred in Phases 2 and 3T. Bromodichloromethane was detected in over 90% of the
systems sampled. The median concentration under the various sample storage conditions ranged
from 5.9 to 14 |ig/L (ppb), and the maximum concentration was 183 |ig/L (ppb). The mean
concentrations of bromodichloromethane in Phases 1, 2, 3T, and 3Q were 18, 18, 17, and 9 |ig/L
(ppb), respectively. Dibromochloromethane was detected in 73% of the systems sampled. The
median concentration ranged from below the detection limit to 3 |ig/L (ppb), and the maximum
value was 280 |ig/L (ppb). The mean concentrations of dibromochloromethane in Phases 1, 2,
3T, and 3Q were 8, 12, 11, and 6 |ig/L (ppb), respectively. The median bromoform concentration
under all sampling conditions was below the detection limit of 0.3 |ig/L (ppb); the maximum value
was 280 |ig/L (ppb). The mean concentrations of bromoform Phases 1, 2, 3T, and 3Q were 3, 4,
4, and 2 |ig/L (ppb), respectively. NOMS was conducted before the promulgation of the total
trihalomethane regulation; therefore, these results may be higher than current levels.
The Community Water Supply Survey (CWSS) was conducted by the EPA in 1978. The
survey examined over 1,100 samples, representing over 450 water supply systems (Brass et al.,
1981). The samples were taken at the treatment plants and in the distribution systems. In the
CWSS, 94% of the surface water supplies and 33% of the ground water supplies were positive
for bromodichloromethane. For surface water supplies, the mean of the positives and the overall
median were 12 and 6.8 |ig/L (ppb), respectively. The mean of the positives for ground water
supplies was 5.8 |ig/L (ppb), and the overall median was below the minimum reporting limit of
0.5 |ig/L (ppb). For dibromochloromethane, 67% of the surface water supplies and 34% of the
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ground water supplies were positive. For surface water supplies, the mean of the positives and
the overall median were 5.0 and 1.5 |ig/L (ppb), respectively. The mean of the positives for
ground water supplies was 6.6 |ig/L (ppb), and the overall median was below the minimum
reporting limit of 0.5 |ig/L (ppb). For bromoform, 13% of the surface water supplies and 26% of
the ground water supplies were positive. The mean concentration of the positives in surface
water supplies was 2.1 |ig/L (ppb), and the overall median was less than 1.0 |ig/L (ppb). The
mean of the positives for ground water supplies was 11 |ig/L (ppb), and the overall median was
below the minimum reporting limit of 0.5 |ig/L (ppb) (Brass et al., 1981).
The Rural Water Survey (RWS) was conducted between 1978 and 1980 by the EPA to
evaluate the status of drinking water in rural America. Samples from over 2,000 households,
representing more than 600 rural water supply systems, were examined. In the RWS, 76% of the
surface water supplies and 13% of the ground water supplies were positive for bromodichloro-
methane, 56% of the surface water supplies and 13% of the ground water supplies were positive
for dibromochloromethane, and 18% of the surface water supplies and 12% of the ground water
supplies were positive for bromoform. For the surface water supplies, the mean of the positives
and the overall median concentrations were 17 |ig/L (ppb) and 11 |ig/L (ppb) for
bromodichloromethane, 8.5 |ig/L (ppb) and 0.8 |ig/L (ppb) for dibromochloromethane, and 8.7
|ig/L (ppb) and <0.5 |ig/L (ppb) for bromoform. For the ground water supplies, the mean of the
positives was 7.7 |ig/L (ppb) for bromodichloromethane, 9.9 |ig/L (ppb) for dibromochloro-
methane, and 12 |ig/L (ppb) for bromoform. The overall median for ground water supplies was
below the minimum reporting limit of 0.5 |ig/L (ppb) for all three brominated trihalomethanes
(Brass, 1981).
The Ground Water Supply Survey (GWSS) was conducted from December 1980 to
December 1981 by the EPA to develop data on the occurrence of volatile organic chemicals in
ground water supplies. Out of a total of 945 ground water systems that were sampled, 466
systems were chosen at random, and the remaining 479 systems were chosen on the basis of
location near industrial, commercial, and waste disposal activities. Samples were collected at or
near the entry to the distribution system, and trihalomethane formation was allowed to continue
without quenching after sample collection. For bromodichloromethane, the median of the
positives for the randomly chosen systems serving greater than 10,000 people was 1.4 |ig/L (ppb),
and the occurrence rate was 36%. For the randomly chosen smaller systems, the median positive
concentration was 1.6 |ig/L (ppb), and the occurrence rate was 54%. The nonrandomly chosen
systems had a median positive concentration of 2.1 |ig/L (ppb) and an occurrence rate of 51%.
For dibromochloromethane, the median positive concentration and the occurrence rate for the
randomly chosen systems serving greater than 10,000 people were 2.1 |ig/L (ppb) and 31%,
respectively; these values for the smaller systems were 2.9 |ig/L (ppb) and 52%. The
nonrandomly chosen systems had a median positive concentration of 3.9 |ig/L (ppb) and an
occurrence rate of 46%. For bromoform, the median positive concentration was 2.4 |ig/L (ppb)
for the randomly chosen systems serving greater than 10,000 and 3.8 |ig/L (ppb) for the randomly
chosen systems serving fewer than 10,000 people, with occurrence rates of 16% and 31%,
respectively. The nonrandomly chosen systems had a median positive concentration of 4.2 |ig/L
(ppb) and an occurrence rate of 31% (Westrick et al., 1983).
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The National Screening Program for Organics in Drinking Water (NSP), sponsored by the
EPA, was conducted from June 1977 to March 1981 and sampled 169 systems nationwide.
Samples were collected at the treatment facilities. For dibromochloromethane, the mean and
median for 130 positives were 17.2 and 10 |ig/L (ppb), respectively. The maximum concentration
found was 131 |ig/L (ppb) (Boland, 1981).
The Technical Support Center (TSC) of the Office of Ground Water and Drinking Water
(OGWDW) maintains a ground water contaminant database. For both bromodichloromethane
and dibromochloromethane, the database contains 4,439 samples taken at the treatment facilities
from nineteen states between 1984 and 1991. For bromodichloromethane, the mean
concentration was 5.6 |ig/L (ppb), and the median was 3 |ig/L (ppb). For dibromochloromethane,
the mean concentration was 3.0 |ig/L (ppb), and the median was 1.7 |ig/L (ppb). For bromoform,
the database contains 1409 samples from 19 states taken at treatment facilities between 1984 and
1991. The mean and median concentrations were determined to be 2.5 |ig/L (ppb) and 1 |ig/L
(ppb), respectively (U.S. EPA, 1991).
Thirty-five water utilities nationwide, 10 of which were located in California, were
sampled for bromodichloromethane, dibromochloromethane, and bromoform in the clearwell
effluent. Samples were taken for four quarters (spring, summer, and fall in 1988 and winter in
1989). The median bromodichloromethane concentration for all four quarters was 6.6 |ig/L
(ppb), with the medians of the individual quarters reported as 6.9, 10, 5.5 and 4.1 |ig/L (ppb),
respectively, and with a maximum value of 82 |ig/L (ppb). For all four quarters, 75% of the
measured concentrations were less than 14 |ig/L (ppb). The median dibromochloromethane
concentration for all four quarters was 3.6 |ig/L (ppb), with the medians of the individual quarters
reported as 2.6, 4.5, 3.8 and 2.7 |ig/L (ppb), respectively, and with a maximum value of 63 |ig/L
(ppb). For all four quarters, 75% of the data were below 9.1 |ig/L (ppb). The median bromoform
concentration for all four quarters was 0.57 |ig/L (ppb), with the medians of the individual
quarters reported as 0.33, 0.57, 0.88, and 0.51 |ig/L (ppb), respectively, and with a maximum
value of 72 |ig/L (ppb). For all four quarters, 75% of the bromoform concentrations were below
2.8 |ig/L (ppb) (Krasner et al., 1989; U.S. EPA 1989a; 1989b).
The EPA's Technical Support Center compiled a database from its disinfection
by-products field studies. The studies included a chlorination by-products survey, conducted
from October 1987 to March 1989. In this survey, concentrations of bromodichloromethane,
dibromochloromethane, and bromoform were determined in finished water from the treatment
plant and in the distribution system. Systems using both surface water sources and ground water
sources were analyzed.
Mean concentrations of bromodichloromethane, dibromochloromethane, and bromoform
in finished water at the treatment plants were determined for surface water systems serving both
greater than and less than 10,000 people. Forty-two samples were taken from systems serving
more than 10,000 people, and 20 samples were taken from systems serving fewer than 10,000
people. The mean concentration of bromodichloromethane was 12.7 |ig/L (ppb) in samples from
systems serving more than 10,000 people (90th percentile, 25.0 |ig/L (ppb)) and 17.0 |ig/L (ppb)
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for samples from the smaller systems (90th percentile, 29.5 |ig/L (ppb)). The mean
dibromochloromethane concentrations was 4.7 |ig/L (ppb) for samples from the larger systems
(90th percentile, 13.8 |ig/L (ppb)) and 6.9 |ig/L (ppb) for samples from the smaller systems (90th
percentile, 24.2 |ig/L (ppb)). The mean concentrations for bromoform were 0.7 |ig/L (ppb) (90th
percentile, 1.5 |ig/L (ppb)) and 0.9 |ig/L (ppb) (90th percentile, 4.9 |ig/L (ppb)) in samples from
the larger systems and samples from the smaller systems, respectively (U.S. EPA, 1992a).
Mean bromodichloromethane, dibromochloromethane, and bromoform concentrations in
distribution systems of these surface water systems also were analyzed. Thirty-nine samples were
taken from systems serving more than 10,000 people, and 11 samples were from systems serving
fewer than 10,000 people. The mean bromodichloromethane concentrations in the larger systems
and the smaller systems were 17.4 |ig/L (ppb) (90th percentile, 35.3 |ig/L (ppb)) and 24.8 |ig/L
(ppb) (90th percentile, 51.0 |ig/L (ppb)), respectively. The mean dibromochloromethane
concentrations were 6.3 |ig/L (ppb) (90th percentile, 17.3 |ig/L (ppb)) and 10.4 |ig/L (ppb) (90th
percentile, 35.0 |ig/L (ppb)), respectively. Mean bromoform concentrations were 0.8 |ig/L (ppb)
(90th percentile, 3.1 |ig/L (ppb)) and 1.4 |ig/L (ppb) (90th percentile, 5.1 |ig/L (ppb)),
respectively (U.S. EPA, 1992a).
Ground water systems serving less than 10,000 people were analyzed for
bromodichloromethane, dibromochloromethane, and bromoform in seven finished water samples
and in five distribution system samples. Mean bromodichloromethane concentrations in the
finished water samples and in the distribution system samples were 1.1 |ig/L (ppb) (90th
percentile, 2.6 |ig/L (ppb)) and 2.2 |ig/L (ppb) (90th percentile, 5.4 |ig/L (ppb)), respectively.
Mean dibromochloromethane concentrations were 0.6 |ig/L (ppb) (90th percentile, 1.0 |ig/L
(ppb)) and 1.8 |ig/L (ppb) (90th percentile, 3.6 |ig/L (ppb)), respectively. Mean bromoform
concentrations were 0.6 |ig/L (ppb) (90th percentile, 2.6 |ig/L (ppb)) and 2.3 |ig/L (ppb) (90th
percentile, 10 |ig/L (ppb)), respectively.
For ground water systems serving more than 10,000 people, dibromochloromethane and
bromoform were not detected in single samples taken at the plant and from the distribution
system, based on a detection limit of 0.2 |ig/L (ppb). Bromodichloromethane concentrations in
the plant and distribution system samples were 0.2 and 0.4 |ig/L (ppb), respectively (U.S. EPA,
1992a).
The U.S. Geological Survey conducted an assessment of volatile organic compounds in
untreated ambient groundwater of the conterminous United States based on samples collected
between 1985 and 1995 from 2948 wells. The sampled wells were located in rural and urban
areas and included wells used for drinking and non-drinking water purposes. A minimum
reporting level of 0.2 //g/L (ppb) was used for most of the compounds, including
bromodichloromethane, dibromochloromethane, and bromoform. In samples from the 406 urban
wells assessed, bromodichloromethane, dibromochloromethane, and bromoform were detected in
3.0%, 2.8%, and 2.8% of the wells examined, respectively. In samples from the 2542 rural wells
examined, these compounds were detected in 0.8%, 0.6%, and 0.4% of the wells, respectively.
The measured concentration of the compounds in well water were reported in summary graphics
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February 20, 2003
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only. Thus, the values reported here are approximate based on visual inspection of the figures.
The median concentrations measured in the positive samples from the urban wells was
approximately 1.0 //g/L (ppb) for all three compounds, while the maximum concentrations of
bromodichloro-methane, dibromochloromethane, and bromoform in the urban wells were
approximately 11, 11, and 13 /i-g/L (ppb), respectively. The median concentrations measured in
the positive samples from the rural wells were approximately 0.4 to 0.5 //g/L (ppb) for all three
compounds, while the maximum concentrations of bromodichloromethane,
dibromochloromethane, and bromoform in the rural wells were approximately 7, 10, and 18 //g/L
(ppb), respectively.
The most recent survey of the occurrence of brominated trihalomethanes in public water
supplies (PWSs) serving at least 100,000 persons resulted from the Information Collection Rule
(ICR) promulgated in May of 1996 for disinfectants and disinfection byproducts (D/DBPs). The
rule covered both surface and ground water systems. Monitoring data were collected from about
300 water systems operating 501 plants over thel8-month period between July 1997 and
December 1998. At each plant, samples were collected monthly and analyzed for a variety of
D/DBPs on a monthly or quarterly basis. Bromodichloromethane, dibromochloromethane, and
bromoform were among the analytes evaluated quarterly (U.S. EPA, 2001a). Five samples were
taken each quarter at each plant - one of the finished water and four of the water in the
distribution system. Of the four samples from the distribution system, one represented a sample
with the same residence time as a finished water sample held for a specific period of time, two
represented approximate average water residence times in the system, and one sample was taken
where water residence time in the system is the longest. For each plant and reporting period,
EPA compiled several summary statistics. The Distribution System (DS) Average value is the
average of the four distribution system samples. The DS High Value is the highest concentration
of the four distribution system samples collected by a plant in a given quarter. The DS High
Value might be from any of the four samples and could vary from quarter to quarter depending on
which sample yielded the highest concentrations in each quarter (U.S. EPA, 2001a). Table IV-1
summarizes the results of all six of the quarterly reporting periods.
U.S. EPA set a minimum reporting level (MRL) for bromodichloromethane,
dibromochloromethane, and bromoform of 1.0 //g/L for the ICR. The MRL is a level below
which systems were not required to report their monitoring results, even if there were detectable
results. Values below the MRL were assigned a value of zero for the purpose of calculating
averages; this assignment affects the calculation of mean values for finished water and DS high
results and calculation of all DS average values.
Recent data for concentrations of brominated trihalomethanes are now available for 117
small surface water plants (serving <10,000 people) from the National Rural Water Association
Survey (NWRA) (U.S. EPA 2001b). Most, but not all, plants that participated in the survey took
two samples at each of three sampling locations. One sample was taken between November,
1999, and March, 2000, and the other between July and November, 2000, for a total of 217 THM
samples. The samples were taken at the finished water location, distribution system average
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February 20, 2003
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residence time location, and maximum residence time location. These data are summarized in
Table IV-2 below.
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February 20, 2003
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Table IV-1. Brominated Trihalomethane Concentrations Measured in U.S. Public
Drinking Water Systems Serving 100,000 or More Persons
Source
Data Type
(a)
Number of
Samples
Median (b)
Mean (b)
90th
Percentile
Range
Bromodichloromethane (/ig/L)
Suface
Water
Finished
1856
6.6
8.2
17.5
<1.0-49
DS Average
1656
8.6
10.2
20.3
0-65.8
DS High
1656
9.9
11.9
23.3
<1.0-73
Ground
Water
Finished
604
< 1.0
7.9
6.80
<1.0-27
DS Average
603
1.80
4.06
11.2
0-35.3
DS High
603
2.8
5.78
16.0
<1.0 - 110
Dibromochloromethane (/ig/L)
Surface
Water
Finished
1853
1.9
4.03
12.0
<1.0-55.1
DS Average
1655
2.40
4.72
13.2
0-67.3
DS High
1655
2.9
5.57
15.0
< 1.0-67.3
Ground
Water
Finished
604
< 1.0
1.38
4.10
<1.0-33
DS Average
602
1.35
3.09
8.94
0-37.5
DS High
602
2.1
4.60
12.9
<1.0 - 85
Bromoform (jWg/L)
Surface
Water
Finished
1853
<1.0
0.998
2.88
<1.0-34
DS Average
1653
0
1.18
3.10
0-34.3
DS High
1653
<1.0
1.48
3.90
<1.0-40
Ground
Water
Finished
602
<1.0
0.838
2.20
<1.0-21
DS Average
599
0.325
1.92
4.78
0-28.8
DS High
599
1.2
2.95
7.72
<1.0-391
(a) Finished = sample location after treatment, before entering the distribution system (DS); DS Average =
average of four sample locations in the DS; DS High = the highest concentration of the four distribution system
samples collected by a plant in a given quarter. For purposes of calculations, all values below the minimum
reporting level (MRL) of 1.0 .vg/L for all three compounds were assigned a value of zero.
(b) Median and mean of all samples including those below the MRL.
Source: Disinfectants and Disinfection Byproducts (D/DBPs) ICR Data, U.S. EPA (2001a).
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Table IV-2 NRWA Brominated Trihalomethane Results for Small Surface Water Plants
THM
Sample
Location
Mean
Median
90th
Percentile
Range
Bromodichloromethane
(Mg/L)
Finished
11.2
6.5
26.6
0 - 84.4
DS Average
14.3
9.4
32.2
0 - 100.3
DS Max
15.9
10.2
34.2
0 - 121.1
Dibromochloromethane
(Mg/L)
Finished
5.0
1.1
13.4
0 - 83.1
DS Average
6.1
1.5
16.3
0 - 99.0
DS Max
6.7
1.9
17.1
0 - 91.6
Bromoform (//g/L)
Finished
4.0
0
1.2
0 - 333.4
DS Average
4.6
0
1.2
0 - 340.5
DS Max
4.5
0
1.3
0 - 349.7
Median and mean of all samples, including those below the detection limit.
Source: National Rural Water Association Survey.
2. Other Studies
Several less comprehensive surveys have analyzed drinking water for one or more of the
brominated trihalomethanes. An overview of these studies is provided below.
The EPA Region V Organics Survey sampled finished water from 83 sites in a region that
includes Illinois, Indiana, Michigan, Minnesota, Ohio, and Wisconsin. Bromoform was found at a
median concentration of the positives of 1 |ig/L (ppb) and a maximum level of 7 |ig/L (ppb). A
total of 14% of the locations sampled contained detectable levels of bromoform (U.S. EPA,
1980). Kelley (1985) surveyed 18 drinking water plants in Iowa for trihalomethanes, and
detected bromoform in five water supplies at concentrations ranging from 1.0 to 10 |ig/L (ppb).
The EPA's Five-year Total Exposure Assessment Methodology (TEAM) study measured
the personal exposures of a probability-based sample of residents in several U.S. cities to various
organic chemicals in air and drinking water between 1981 and 1987. As part of the study, running
tap water samples were collected from residences of nearly 850 study participants during the
morning and the evening to test for brominated trihalomethane concentrations. The samples were
quenched with sodium thiosulfate at the time of collection. Tables IV-3, IV-4, and IV-5 show
bromodichloromethane, dibromochloromethane, and bromoform concentrations found in drinking
water from the six cities surveyed. Samples of water were taken from each participating
residence at the household taps and sodium thiosulfate added as a quenching agent.
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Table IV-3 Bromodichloromethane Concentrations in Drinking Water from the U.S. EPA
TEAM Study (jug/L)
Location
Date
Sampled
Sample
Size
%
Meas-
ured
Concentration jug/L (ppb)
Mean
Med-
ian
Max
Percentiles
25%
75%
90%
95%
Elizabeth/
Bayonne,
New Jersey
Fall 1981
340
99.7
13.6
13
23
--
15
16
18
Summer
1982
156
99.8
13.6
12
54
--
15
18
20
Winter
1983
49
100
5.4
5.8
16
--
7.1
8.3
8.3
Los Angeles,
California
Winter
1984
117
93
11
12
23
5.1
16
17
20
Summer
1984
52
96
20
24
38
7.7
31
33
37
Winter
1987
9
89
19
24
31
--
--
--
--
Summer
1987
7
100
26
27
36
--
--
--
--
Antioch/
Pittsburg,
California
Spring
1984
71
96
21
17
47
2.4
36
45
47
Devils Lake,
North
Dakota
Fall 1982
24
73
0.21
0.18
1.0
--
--
--
--
Greensboro,
North
Dakota
Fall 1982
24
93
7.1
7.8
11
--
9.2
--
--
Baltimore,
Maryland
Spring
1987
10
100
10
10
13
--
--
--
--
Adopted from Hartwell, (1987), Wallace et al. (1987), Wallace et al. (1988), and Wallace (1992) by U.S. EPA
(1994b). Mean and median values of all samples, including those below the quantitation limit.
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Table IV-4 Dibromochloromethane Concentrations in Drinking Water from the U.S. EPA
TEAM Study
Location
Date
Sampled
Sample
Size
%
Meas-
ured
Concentration jug/L (ppb)
Mean
Med-
ian
Max
Percentile
25%
75%
90%
95%
Elizabeth/
Bayonne,
New Jersey
Fall 1981
340
99.7
2.4
2.4
8.4
--
2.7
3.2
3.4
Summer
1982
156
99.8
2.1
1.9
7.2
--
2.4
3.1
3.8
Winter
1983
49
93
1.4
1.6
3.0
--
1.8
2.0
2.1
Los Angeles,
California
Winter
1984
117
89
9.4
11
19
2.4
15
17
18
Summer
1984
52
85
28
32
55
15
42
43
48
Winter
1987
9
89
10
12
17
--
--
--
--
Summer
1987
7
100
24.7
18
70
--
--
--
--
Antioch/
Pittsburg,
California
Spring
1984
71
85
8
6.4
21
0.98
15
18
19
Devils Lake,
North Dakota
Fall 1982
24
18
0.09
0.06
0.45
--
0.06
--
--
Greensboro,
North Dakota
Fall 1982
24
93
1.2
1.2
1.9
--
1.5
--
--
Baltimore,
Maryland
Spring
1987
10
100
2.7
2.6
3.5
--
--
--
--
Adopted from Hartwell, (1987), Wallace et al. (1987), Wallace et al. (1988), and Wallace (1992) by U.S. EPA
(1994b). Mean and median values of all samples, including those below the quantitation limit.
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Table IV-5 Bromoform Concentrations in Drinking Water from the U.S. EPA
TEAM Study
Location
Date
Sampled
Sample
Size
%
Meas-
ured
Concentration jug/L (ppb)
Mean
Med-
ian
Max
Percentile
25%
75%
90%
95%
Los Angeles,
California
Winter
1984
117
69
0.78
0.54
12
0.34
0.92
1.2
1.5
Summer
1984
52
90
8.08
3.0
78
2.0
5.9
13
53
Winter
1987
9
89
3.2
3.2
4.7
--
--
--
--
Summer
1987
7
100
25.5
9.6
113
--
--
--
--
Antioch/
Pittsburg,
California
Spring
1984
71
69
0.78
0.58
2.0
0.19
1.2
1.8
1.9
Adopted from Wallace (1992) by U.S. EPA (1994b). Mean and median values of all samples, including those
below the quantitation limit. Bromoform was measured in fewer than 10% of samples from the other four cities in
the TEAM study and are not presented here
Furlong and D'itri (1986) reported that a survey of water treatment plants in Michigan
detected bromodichloromethane in 35 of 40 plants at a median concentration of 2.7 |ig/L (ppb)
and a maximum of 54.2 |ig/L (ppb); the mean of the positive samples was 7.4 |ig/L (ppb).
Dibromochloromethane was also detected in 30 plants at a median concentration of 2.2
|ig/L (ppb) and a maximum of 39.6 |ig/L (ppb); the mean of the positives was 5.1 |ig/L (ppb).
Bromoform was detected at three of 40 plants sampled at concentrations of 0.9, 1.3, and 1.6 |ig/L
(ppb).
Fair et al. (1988) analyzed drinking water from three community water supplies for
chlorination by-products. Bromodichloromethane concentrations ranged from 7.5 to 30 |ig/L
(ppb) in finished water and from 9.9 to 36 |ig/L (ppb) in the distribution systems.
Dibromochloromethane concentrations ranged from less than 0.5 to 19 |ig/L (ppb) in finished
water at the plant and from less than 0.5 to 23 |ig/L (ppb) in the distribution systems. Bromoform
concentrations ranged from less than 0.5 to 2.5 |ig/L (ppb) in finished water and from less than
0.5 to 3.1 |ig/L (ppb) in the distribution systems.
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Wallace et al. (1982) analyzed tap water for bromodichloromethane as part of a study to
determine individual exposures to volatile organics during normal daily activities of students at the
University of North Carolina, Chapel Hill. Bromodichloromethane was detected in 7 of 7 samples
of tap water, at concentrations ranging from 15 to 20 |ig/L (ppb), with a mean of 17 |ig/L (ppb).
The detection limit was 0.1 |ig/L (ppb).
Chang and Singer (1984) analyzed the bromoform concentration in drinking water
samples prepared by the desalination of seawater. After pretreatment using either activated
carbon or ultrafiltration, but prior to reverse osmosis treatment, bromoform concentrations were
13 ± 14 and 110 ± 59 |ig/L (ppb), respectively. After reverse osmosis was completed, the
finished water product contained bromoform concentrations ranging from 2.0 to 51 |ig/L (ppb)
(mean, 20.17 |ig/L (ppb)) when activated carbon was used as a pretreatment and 127 |ig/L (ppb)
when ultrafiltration was used. In the reverse osmosis treatment, three reverse osmosis membranes
were evaluated. The cellulose triacetate filter resulted in concentrations of 51 |ig/L (ppb), while
the polyether/urea thin film spiral wound membrane and the polysulfone membrane filters which
resulted in final concentrations of 5.0 |ig/L (ppb) and 2.25 |ig/L (ppb), respectively.
Bromodichloromethane, dibromochloromethane, and bromoform were detected in 9.5 to
12.8% of drinking water samples collected in 1987 in Nassau County, New York. The county
draws its drinking water from underground aquifers. Bromodichloromethane and
dibromochloromethane had similar concentration profiles, being detected in approximately 10%
and 8.5%) of the samples, respectively, at concentrations less than 4.9 ppb. The detection limit
was 1 ppb for each chemical. Bromoform was detected in 8%> of the samples at 2 to 4.9 ppb, in
2.5% of the samples at 5 to 9.9 ppb, and in less than 1%> of the samples at 10 to 49.9 ppb. The
detection limit was 2 ppb. None of the drinking water samples contained more than 50 ppb of any
of the trihalomethanes, and less than 1%> of the samples contained between 10 and 49.9 ppb of the
brominated compounds (Moon et al. 1990).
U.S. EPA conducted a study of contaminants in household water in nine residences as part
of a larger study of health risks due to environmental contamination in the Lower Rio Grande
Valley (Berry et al., 1997). Samples of water used for drinking were taken once during a 3-day
period in the spring and once during a 2-day period in the summer of 1993. Water used for
drinking in the nine residences could be traced to one of three sources: the municipal water supply
of Brownsville, Texas, vended water supplies (municipal water that had undergone further
treatment), and well water. Samples were collected using U.S. EPA protocols, including quality
assurance samples and field blanks. The detection and minimum quantitation limits for each
analyte were documented in other reports. Bromodichloromethane, dibromochloromethane, and
bromoform were detected in the household water of seven of the nine residences during the spring
and in five of the nine residences during the summer (Berry et al., 1997). During the spring, the
minimum, median, and maximum concentrations of bromodichloromethane for the seven positive
samples were 3.2, 5.2, and 24.4 ju-g/L (ppb), respectively. For dibromochloromethane, the values
were 3.3, 5.1, and 17.3 jug/L (ppb), respectively. For bromoform, the values were 1.0, 3.0, and
14.1 jug/L (ppb), respectively. During the summer, the minimum, median, and maximum
concentrations of bromodichloromethane in the five positive samples were 2.3, 7.7, and 34.3 ju.g/L
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(ppb), respectively. For dibromochloromethane, the values were 1.8, 7.6, and 49.9 //g/L (ppb),
respectively, and for bromoform, the values were 1.6, 7.8, and 31.7 //g/L (ppb), respectively.
Weisel et al. (1999) examined concentrations of trihalomethanes in the tap water of the
homes of 49 women in New Jersey. The 49 residences were selected so that approximately half
would represent the lower extreme of trihalomethane contamination and half the upper extreme of
trihalomethane contamination identified in a previous study. Samples were stored unquenched on
ice after collection and were analyzed within 24 hours. The three brominated trihalomethanes
were detected in all 49 samples. The mean (± standard deviation) concentrations of
bromodichloromethane, dibromochloromethane, and bromoform were 5.7 ± 8.6, 2.0 ± 2.1, and
0.73 ± 0.90 /ig/L (ppb), respectively. The median values for the three compounds were 2.6, 1.4,
and 0.45 //g/L (ppb), respectively. These values are not representative of New Jersey, because of
the selection criteria for the residences. The ranges (minimum to maximum) of concentrations
measured for each compound were 0.06 to 48 //g/L (ppb) for bromodichloromethane, 0.14 to 9.7
Mg/L (ppb) for dibromochloromethane, and 0.03 to 4.21 /ig/L (ppb) for bromoform.
3. Estimates of Tap Water Ingestion Exposure to Brominated Trihalomethanes
a. Estimates Based on ICR Data for Disinfection Byproducts
The data from EPA's ICR for disinfectants and disinfection byproducts (U.S. EPA 2001a)
offer several advantages over the other national studies for purposes of estimating national
exposure levels of adults in the United States to brominated trihalomethanes via ingestion of
drinking water. First, they are recent and reflect relatively current conditions. Second, data of
very similar quality and quantity were collected systematically from a large number of plants (501)
and systems (approximately 300), including both surface and ground water systems. Third, the
mean, median, and 90th percentile value were estimated on the basis of all samples taken, not just
the sample detects. Thus, these descriptive statistics are representative of the exposures of the
entire populations served by those systems, not just the populations served by systems with higher
concentrations of these compounds. However, this study can not be considered representative of
smaller public water supplies or water supplies from the most highly industrialized or
contaminated areas.
Table IV-6 presents estimated drinking water exposures to brominated trihalomethanes of
the adult populations served by large public water systems (serving 100,000 or more persons)
based on the ICR Occurrence Data (U.S. EPA, 2001a). Exposure was calculated by multiplying
the concentration of individual brominated trihalomethanes in drinking water by the average daily
intake, assuming that each individual consumes two liters of water per day. The annual median,
mean, and upper 90th percentile values are presented for both surface and ground water systems.
Assuming that the DS High value actually represents the average exposure level of persons served
by one plant distribution pipe with the longest water-residence time, the DS High value might be
used to estimate a high-end exposure level. Thus, the 90th percentile of the DS High values are
also presented in Table IV-6.
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Table IV-6 Estimated Drinking Water Exposures to Brominated Trihalomethanes in U.S.
Public Drinking Water Systems Serving More than 100,000 Persons3
Source
Medianb
Meanb
90th Percentileb
DS High 90th
Percentile0
Bromodichloromethane (/ig/person/day)
Surface Water
17
20
40
47
Ground Water
3.6
8.1
22
32
Dibromochloromethane (/ig/person/day)
Surface Water
4.8
9.4
26
30
Ground Water
2.7
6.2
18
26
Bromoform (jwg/person/day)
Surface Water
0
2.4
6.2
7.8
Ground Water
0.65
3.8
9.6
15
a Source: U.S. EPA (2001a). Assumes that each individual consumes 2 liters of water daily. Also assumes that
concentrations at the drinking water tap are similar to concentrations in the distribution system (DS) sampled at
locations considered to be representative of average (DS Average) and highest (DS High) retention times (see
Table IV-1).
b Based on concentrations from the DS Average values.
c Based on the 90th percentile of the DS High values to represent a plausible high-end exposure level.
For bromodichloromethane, the median, mean, and 90th percentile population exposures
from surface water systems are estimated to be 17, 20, and 40 |ig/person/day, respectively. The
same values for populations exposed to bromodichloromethane from ground water systems are
lower - 3.6, 8.1, and 22 |ig/person/day, respectively. For dibromochloromethane, the median,
mean, and 90th percentile population exposures from surface water systems are estimated to be
4.8, 9.4, and 26 |ig/person/day, respectively. The corresponding values for populations exposed
to dibromochloromethane from groundwater system are lower - 2.7, 6.2, and 18 |ig/person/day,
respectively. For bromoform, the median, mean, and 90th percentile population exposures from
surface water systems are estimated to be near 0, 2.4, and 6.2 |ig/person/day, respectively. The
same values for populations exposed to bromoform from ground water systems are higher - 0.65,
3.8, and 9.6 |ig/person/day, respectively.
Average daily intake of dibromochloromethane was also evaluated for determination of
the Relative Source Concentration. The details of this evaluation are presented in Appendix C.
Intake for ingestion was calculated using mean intake rates of 1.2 or 0.6 L/day for total and direct
intake (NRC, 1999), respectively. Direct intake includes consumption of water directly from the
tap, but does not include intake of tap water used for preparation of heated items such tea, coffee,
or soup. Based on the ICR distribution system average concentration of 4.72 |ig/L for
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dibromochloromethane in surface water, the average daily total and direct and ingestion intakes
would be 5.7 and 2.8 |ig/day, respectively. Absorption of dibromochloromethane from tap water
was estimated using methodology described in U.S. EPA (1992c), as modified by Vecchia and
Bunge (2002). The average dermal uptake of dibromochloromethane was estimated to be 2 |ig
per shower or bathing event. Intake via inhalation of dibromochloromethane volatilized during
household activities (e.g., showering, bathing, dishwashing, toilet flushing, etc.) was estimated
using a three-compartment model based on McKone (1987). This model estimated an average
daily inhalation exposure of 7 |ig/day for the volatilized compound. Parallel calculations were not
performed for bromodichloromethane or bromoform, because these compounds are probable
carcinogens. Therefore, in accordance with U.S. EPA policy, RSC analysis was not conducted.
b. Estimates of Ingestion Exposure Based on Other National Studies
Exposure to bromodichloromethane, dibromochloromethane, and bromoform in drinking
water from ground water supplies can be estimated from the median levels found in the GWSS.
Based on the range of median levels (1.4-2.1 |ig/L (ppb)) and a consumption rate of two liters
per day, the median exposure to bromodichloromethane may range from 2.8 to 4.2 |ig/day.
Similarly, median exposure to dibromochloromethane may range from 4.2 to 7.8 |ig/day, and for
bromoform, median exposure may range from 4.8 to 8.4 |ig/day. Exposure to
bromodichloromethane from surface water supplies can be estimated based on the range of
median values observed under different conditions in NOMS, which mainly sampled surface water
systems. Based on a range of 5.9-14 |ig/L (ppb), exposure to bromodichloromethane from
surface water is estimated to be between 12 and 28 |ig/day. Similarly, based on the range of
medians reported for dibromochloromethane concentrations, the median exposure is estimated to
be up to 6 |ig/day. The median levels of bromoform in the surface water supplies have been
found to be less than the EPA Drinking Water minimum reporting levels (MRLs) of 0.5-1 |ig/L
(ppb). An estimate of exposure based on the MRLs will be overly conservative because the actual
concentration of bromoform is not detectable. Based on the range of MRLs, 0.5-1 |ig/L (ppb),
the exposure to bromoform is estimated to range from 1 to 2 |ig/day for surface water supplies.
Ingestion exposure to brominated trihalomethanes in drinking water can also be estimated
from the concentrations found at the tap in the TEAM studies. Table IV-7 presents median,
mean, 90th percentile, and 95th percentile estimates of daily intakes of bromodichloromethane,
dibromochloromethane, and bromoform, based on an assumed drinking water ingestion rate of 2
liter per day. Table IV-7 provides estimates for those locations and seasons with a sample size of
at least 50, with one exception. Devils Lake, ND, with a samples size of only 24, is added to
represent an area with low concentrations. Thus, the influence of small sample size on
distributional statistics should be minimized in Table IV-7. The median, mean, and 90th percentile
values in Table IV-7 for the TEAM study can be compared with the corresponding values in
Table IV-6 for the ICR Occurrence data.
Table IV-7 demonstrates that concentrations of brominated trihalomethanes are lower in
winter than in summer, as would be expected on the basis of temperature. In this sample of
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Table IV-7. Estimated Distribution of Drinking Water Exposures to Brominated
Trihalomethanes for Populations in U.S. EPA TEAM Study (a)
Location
Season
Year
Median (b)
Mean (b)
90th
Percentile (b)
95th
Percentile
Bromodichloromethane (/ig/person/day)
Elizabeth/Bayone NJ
summer 82
24
27
36
40
winter 83
12
11
17
17
Los Angeles, CA
summer 84
48
40
66
74
winter 84
24
22
34
40
Antioch/Pittsburg, CA
spring 84
34
42
90
94
Devils Lake, ND
fall 82
0.36
0.42
<2.0
<2.0
Dibromochloromethane (/ig/person/day)
Elizabeth/Bayone NJ
summer 82
3.8
4.2
6.2
7.6
winter 83
3.2
2.8
4.0
4.2
Los Angeles, CA
summer 84
64
56
86
96
winter 84
22
19
34
36
Antioch/Pittsburg, CA
spring 84
13
16
36
38
Devils Lake, ND
fall 82
0.12
0.2
<0.9
<0.9
Bromoform (jwg/person/day)
Los Angeles, CA
summer 84
6.0
16.2
26
100
winter 84
1.1
1.6
2.4
3.0
Antioch/Pittsburg, CA
spring 84
1.2
1.6
3.6
3.8
(a) Intakes estimated from data in Tables IV-3, IV-4, and IV-5 assuming a water ingestion rate of 2 liters per
day. Selected locations and seasons with samples sizes over 50. Added Devils Lake, ND, to represent an area
with low air concentrations.
(b) Median, mean, and upper percentiles estimated for entire population of city.
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geographic locations, estimates of the average of the population intakes of bromodichloromethane
from drinking water range from 0.42 to 42 //g/person/day. The upper 90th percentile estimates
range from <2.0 to 90 /ig/person/day. Estimates of the average population intake of
dibromochloromethane from drinking water range from 0.2 to 56 /ig/person/day. The upper 90th
percentile estimates range from < 0.9 to 86 //g/person/day. Estimates of the average of the
population intakes of bromoform, for those areas in which bromoform was measurable in a
majority of the samples, range from 1.6 to 16.2 //g/person/day. The upper 90th percentile
estimates range from 2.4 to 26 //g/person/day. Four of the six locations
in the TEAM study, however, had a low frequency (less than 10%) of detection of bromoform in
measurable quantities.
c. Sources of Uncertainty in Estimates of Exposure from Drinking Water
Sources of uncertainty in the estimates of ingestion exposure include use of different
analytical methods, failure to report quantitation limits, using measures near the detection limit,
failure to report how nondetects are handled when averaging values (e.g., set to zero or one half
the detection limit), and failure to report sample storage method and duration. In addition, many
environmental factors influence the concentrations of these compounds in drinking water at the
tap and in vended or bottled waters used for drinking. These factors include season and
temperature, geographic location, source of water, residence time in distribution system, and
others.
B. Exposure from Sources Other Than Drinking Water
1. Dietary Intake
a. Measured Concentrations in Foods and Beverages
Information on the levels of brominated trihalomethanes in foods and beverages is limited.
Chlorine is used in food production for applications such as the disinfection of chicken in poultry
plants and the superchlorination of water at soda and beer bottling plants (Borum, 1991).
Therefore, the possibility exists for contamination of foodstuffs by disinfection by-products with
resulting dietary exposure. The occurrence of bromodichloromethane in foods and beverages is
the best characterized of the three compounds. Less information is available concerning the
occurrence of dibromochloromethane or bromoform in foods and beverages in the United States.
Some information is available from international studies, but may not be relevant to U.S.
occurrence because of different water treatment and food processing practices. The available
U.S. and international studies are summarized below.
Entz et al. (1982) analyzed food samples from Elizabeth, NJ, Chapel Hill, NC, and
Washington, DC. for bromodichloromethane. A total of 39 different food items from each city
were collected according to standards set for the FDA's Total Diet Market Basket Study. The
Adult Market Basket, representing the diet of a teenage male, is divided into 12 food groups.
Individual foods are prepared as generally consumed in the home and foods from each group are
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blended together in "the proper proportions" to form composites. In this study, foods were
blended into four composites representing dairy products; meat, fish and poultry; oils, fats and
shortening; and beverages. The estimated limit of quantitation for bromodichloromethane in each
of these composites was 2.3, 4.5, 8.3, and 0.5 ng/g, respectively. Five sets of each composite
were tested for a total of 20 composites. Bromodichloromethane was detected in one dairy
composite at 1.2 ppb and two beverage composites at 0.3 ppb and 0.6 ppb. Analysis of individual
foods from the beverage and dairy composites found bromodichloromethane in three samples of
cola soft drinks at concentrations of 2.3 ppb, 3.4 ppb, and 3.8 ppb and in one sample of butter at
7 ppb.
Uhler and Diachenko (1987) sampled 38 food and beverage products from 15 food
processing plants in nine states. Plants were chosen on a "worst-case" basis from areas where
contaminated water would most likely be used in processing. In addition, processing plants were
chosen for study only if they produced high fat content food that came in contact with water
during processing or contained a high percentage of added water. Samples containing less than 1
ng/g were considered nondetects. Bromodichloromethane was detected in 6 out of 37 tested food
tested at the following levels: two samples of clear sodas at 1.2 and 2.3 ng/g (ppb) and one
sample of dark cola at 1.2 ng/g (ppb) out of fifteen soft drinks, and three of six samples of ice
cream at 0.6 to 2.3 ng/g (ppb). Bromodichloromethane was not found in any of the eight cheese
samples analyzed.
U.S. EPA (1985) reported that bromodichloromethane was identified in bacon. No
further information on sample size, detection limit, or study methodology was provided.
Abdel-Rahman (1982) analyzed various soft drinks for bromodichloromethane and found
average levels ranging from 0.2 to 6.6 |ig/L (ppb) for colas and from 0.1 to 0.2 |ig/L (ppb) for
clear soft drinks (Abdel-Rahman, 1982). In Italy, Cocchioni et al. (1996) analyzed 61 samples of
different commercially prepared beverages and 94 samples of mineral waters for volatile organo-
halogenated compounds. In the prepared beverages, they found maximum concentrations of
bromodichloromethane, dibromochloromethane, and bromoform of 40.6, 13.9, and 10.7 |ig/L
(ppb), respectively. The frequencies of detection of these three compounds in prepared beverages
were 46% (28/61), 43% (26/61), and 11% (7/61), respectively, with detection limits for all three
compounds of less than 1 |ig/L (ppb). In contrast, the maximum concentration of any of the
halogenated organic compounds identified in mineral water, including chloroform, was 5.79 |ig/L
(ppb).
McNeal et al. (1995) examined 27 different prepared beverages and mineral waters in the
United States for bromodichloromethane, dibromochloromethane, and bromoform at detection
limits of 0.1, 0.1, and 0.2 ng/g (ppb), respectively. Bromoform was not detected in any of the
samples. Bromodichloromethane and dibromochloromethane were detected at 12 and 1 ng/g
(ppb), respectively, in only one of seven types of mineral and sparkling waters examined. The
positive sample was the only sparkling and flavored water of the group,. Bromodichloromethane
was found in 1 of 5 flavored noncarbonated beverages examined, a fruit drink, at a concentration
of 5 ng/g (ppb); dibromochloromethane was not detected in any of these five beverages.
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Bromodichloromethane was found in all 13 of the types of carbonated soft drinks examined, at
concentrations ranging from 1 to 4 ng/g (ppb) for 12 of the drinks examined and at 12 ng/g (ppb)
for the thirteenth. Dibromochloromethane was detected in only 4 of the 13 carbonated soft drinks
examined at levels of 0.5 to 2 ng/g (ppb). None of the brominated trihalomethanes was detected
in either of the two types of beer examined.
McNeal et al. (1995) also examined several types of prepared non-beverage foods and
water from canned vegetables in the United States for bromodichloromethane,
dibromochloromethane, and bromoform. None of these compounds was detected in any of the
samples. The foods examined included two types of canned tomato sauce, canned pizza sauce,
canned vegetable juice, vegetable waters from two types of canned green beans and one type of
sweet corn, duck sauces, beef extract, and Lite syrup product.
The U.S. Food and Drug Administration (U.S. FDA, 2000) has analyzed for 18 volatile
organic hydrocarbons (VOCs), including bromodichloromethane and bromoform, in the Total
Diet Study since 1995. Bromodichloromethane and bromoform were analyzed in a subset of 70
food items in 14 Market Baskets. During the period 1995 to 1999, bromodichloromethane was
detected in one sample each of 11 non-beverage food items (sliced bologna, fried eggs, canned
pork and beans, smooth peanut butter, homemade cornbread, raw orange, canned pineapple,
boiled collards, red tomato, green pepper, and fast-food hamburger) (U.S. FDA, 2000). The
detected concentrations ranged from 10 to 16 ppb, with the exception of fast food hamburger
which contained 37 ppb. Bromodichloromethane was detected in one sample of bottled apple
juice at a concentration of 33 ppb. The mean detected concentration of bromodichloromethane in
three samples of tap water was 18 ppb. Dibromochloromethane was not included in the list of
VOC analytes for the Total Diet Study. Bromoform was listed as an analyte, but no detections
were reported in the data summary for 1991 to 1999. The detection limits for
bromodichloromethane and bromoform were not reported.
Imaeda et al. (1994) examined bean curd commercially available in Japan for
trihalomethanes. Neither bromoform nor dibromochloromethane were detected in any of the
samples at a detection limit of 0.1 ppb. Bromodichloromethane was detected in 6 of 10 samples
of bean curd at concentrations ranging from 1.2 to 5.2 ppb and in 1 of 10 samples of the water in
the bean curd packages at 5.2 ppb.
Kroneld and Reunanen (1990) analyzed for brominated trihalomethanes in samples of
pasteurized and unpasteurized cow's milk collected in Turku, Finland. The average concentration
of bromodichloromethane measured in pasteurized milk was 0.008 |ig/L (ppb) (range,
undetectable to 0.03 |ig/L (ppb), detection limit not specified). Dibromochloromethane was
detected in only one sample of pasteurized milk at 5 |ig/L (ppb). Traces of bromoform were
detected but not quantified. Brominated trihalomethanes were not detected in unpasteurized milk.
Their presence in pasteurized milk was considered to result from use of chlorinated water during
processing.
b. Estimated Dietary Intake
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Estimates for dietary intake of brominated trihalomethanes by residents of the United
States were not identified in the materials reviewed for this document. Furthermore, information
on the levels in U.S. foods is too limited to independently calculate a reliable estimate. However,
the available data suggest that the concentrations of brominated trihalomethanes in non-beverage
foods are likely low. The apparently low concentrations of brominated trihalomethanes in non-
beverage foods are consistent with the physical and chemical properties of these compounds. The
levels of individual brominated trihalomethanes in beverages prepared in the United States appear
to be less than or about equal to levels measured in disinfected surface water.
Toyoda et al. (1990) analyzed the dietary intake of bromodichloromethane,
dibromochloromethane, and bromoform for 30 Japanese housewives in Nagoya and Yokohama,
Japan. Duplicate portions of daily meals were collected for three consecutive days and sampled
for all three brominated trihalomethanes. The types of food consumed were not reported. This
omission prevents a meaningful comparison of the studied diet to that consumed by the U.S.
population. The detection limits for bromodichloromethane, dibromochloromethane, and
bromoform were reported to be 0.1, 0.2, and 0.5 ppb, respectively. The concentration of
bromodichloromethane ranged from undetectable to 1.7 ppb (average, 0.3 ± 0.3 ppb SD). The
mean daily intake of bromodichloromethane was estimated to be 0.6 ± 0.5 |ig/day. The
concentration of dibromochloromethane ranged from undetectable to 0.6 ppb (average, 0.1 ±
0.2 ppb), and the mean dietary intake was estimated to be 0.3 ± 0.3 |ig/day. The concentration of
bromoform ranged from undetectable to 8.1 ppb (average, 0.5 ± 1.3 ppb). The mean dietary
intake of bromoform was estimated to be 0.9 ±1.3 |ig/day.
Brominated trihalomethanes have been detected in a number of beverages. In conducting
an exposure assessment, the potential exposures from drinking prepared beverages would not be
added to the default assumption of an adult consuming 2 liters of drinking water per day. Instead,
the prepared beverages would be considered part of the 2 liters of fluid intake per person per day.
2. Air Intake
a. Concentrations in Outdoor Air
Brominated trihalomethanes are usually found in outdoor air at low concentrations when
all data, including nondetects, are considered. Brodzinsky and Singh (1983) reviewed,
summarized, and critically evaluated existing data for brominated trihalomethane concentrations in
ambient outdoor air for several urban/suburban or source dominated locations across the United
States (Table IV-8). No concentration data were available for rural or remote areas. The authors
reported mean, median, first and third quartile values, and minimum and maximum values by city.
In addition, they reported the same measures when the data were grouped by type of location
(i.e., urban/suburban or source dominated), and when all data were combined. Ambient air
concentrations were reported for bromodichloromethane at Magnolia, AR, El Dorado, TX,
Chapel Hill, NC, and Beaumont, TX. Bromodichloromethane was detected at mean
concentrations of 0.76 ppt, 1.40 ppt, 120 ppt, and 180 ppt for those four cities, respectively,
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where ppt is expressed as parts per trillion by volume. Dibromochloromethane was detected in
the air samples from Magnolia, AR, El Dorado, TX, Chapel Hill, NC, Beaumont TX, and Lake
Charles, LA at mean concentrations of 0 ppt, 0.48 ppt, 14 ppt, 14 ppt, and 19 ppt, respectively.
Bromoform was detected in air samples from Magnolia, AR, El Dorado, TX, and Lake Charles,
LA, at concentrations of 1.5 ppt, 0.81 ppt, and 50 ppt, respectively. Air concentration data from
these sites were combined for additional statistical analysis. The study authors indicated that a
value of 0.0 was entered for samples below the detection limit. Mean (± standard deviation)
outdoor air concentrations in urban/suburban and source dominated locations, respectively, were
160 ± 29 ppt and 1.2 ± 0.4 ppt for bromodichloromethane; 15 ± 4 ppt and 0.28 ± 0.67 ppt for
dibromochloromethane; and 50 ± 29 ppt and 1.1 ±2.1 ppt for bromoform. Brodzinsky and Singh
(1983) also calculated overall (grand) means based on data from all sites. Grand mean values for
bromodichloromethane, dibromochloromethane, and bromoform were 110 ppt (n = 26, with one
nondetect), 3.8 ppt (n = 89, with 63 nondetects), and 3.6 ppt (n = 78, with 60 nondetects),
respectively. When expressed on a |ig/m3 basis, the corresponding mean values for
bromodichloromethane, dibromochloromethane, and bromoform are 0.74 |ig/m3, 0.032 |ig/m3,
and 0.037 |ig/m3.
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Table IV-8. Selected Concentration Data for Individual Brominated Trihalomethanes (ppt) in Outdoor Air as Summarized in
Brodzinsky and Singh (1983)a'b
City
n
Nondetects
Mean
(Std dev.)
Median
3rd Quartile
Maximum
Reference
Bromodichloromethane
Individual Sites
Beaumont, TX
11
0
180 (100)
180
180
180
Wallace (1981)
Chapel Hill, NC
6
0
120 (210)
120
120
120
Wallace (1981)
El Dorado, AR
7
1
1.4 (0.35)
1.6
1.6
1.6
Pellizzari and Bunch (1979)
Magnolia, AR
2
0
0.76 (0.0)
0.0
0.0
0.76
Pellizzari and Bunch (1979)
Totals
Urban/Suburban
17
0
160 (29)
180
180
180
-
Source Areas
9
1
1.2 (0.41)
1.6
1.6
1.6
-
Grand totals
26
1
110 (82)
120
180
180
-
Dibromochloromethane
Individual Sites
Beaumont, TX
11
0
14 (0.0)
14
14
14
Wallace (1981)
Chapel Hill, NC
6
0
14 (0.0)
14
14
14
Wallace (1981)
El Dorado, AR
40
35
0.48 (0.82)
0.0
0.82
2.5
Pellizzari et al. (1978)
Lake Charles, LA
4
0
19 (9.6)
21
27
27
Pellizzari (1979)
Magnolia, AR
28
28
0.0 (0.0)
0.0
0.0
0.0
Pellizzari et al. (1978)
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Table IV-8 (cont.)
City
n
Nondetects
Mean
(Std dev.)
Median
3rd Quartile
Maximum
Reference
Dibromochloromethane (cont.)
Totals
Urban/Suburban
21
0
15 (4.2)
14
14
27
-
Source Areas
68
63
0.28 (0.67)
0.0
0.0
2.5
-
Grand Totals
89
63
3.8 (6.7)
0.0
2.5
27
-
Bromoform
Individual Sites
El Dorado, AR
46
35
0.81 (0.95)
0.43
1.3
2.7
Pellizzari et al. (1978)
Pellizzari and Bunch (1979)
Lake Charles, LA
4
0
50 (29)
62
68
71
Pellizzari (1979)
Magnolia, AR
28
25
1.5 (3.2)
0.0
0.29
8.3
Pellizzari et al. (1978)
Totals
Urban/Suburban
4
0
50 (29)
62
68
71
-
Source Areas
74
60
1.1(2.1)
0.0
1.3
8.3
-
Grand Totals
78
60
3.6 (12)
0.0
1.5
71
-
a Includes only data considered to be of adequate, good, or excellent quality by the study authors.
b Concentrations are reported as parts per trillion by volume
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Shikiya et al. (1984) analyzed ambient air samples collected at four urban/industrial
locations in the California South Coast Air Basin from November 1982 to December 1983 for the
presence of halogenated hydrocarbons. Data for bromodichloromethane, dibromochloro-
methane, and bromoform were included in this analysis. The sampling locations were El Monte,
downtown Los Angeles, Dominguez, and Riverside. The air samples were analyzed using gas
chromatography with detection by electron capture. The quantitation limit, defined as a level 10
times greater than the noise level, was 10 ppt by volume for all three brominated trihalomethanes.
The detection limit was defined as three times the noise level. Summary data for each compound
included monthly means and composite means. The monthly means were calculated as the
average of all data at a site that were above the quantitation limit for a single month; samples with
concentrations below the limit of detection were not included in the calculations. The composite
means were calculated as the average value of all data for each compound above the quantitation
limit at each site. Most data in this report were presented graphically. A few additional details
were presented in a short summary statement for each chemical. Thirty-five percent of the
samples had bromodichloromethane levels above the quantitation limit of 10 ppt (0.067 |ig/m3).
Peaks in the concentration of bromodichloromethane were observed at various sites in June and
July, with downtown Los Angeles and Dominguez registering the highest monthly means of
approximately 30 ppt (0.20 |ig/m3). The highest reported concentration was 40 ppt (0.27 |ig/m3).
The highest composite mean of 100 ppt (0.67 |ig/m3) for bromodichloromethane was observed at
El Monte. In comparison, the remaining three locations had a composite mean of 20 ppt (0.08
|ig/m3). For dibromochloromethane, only seventeen percent of the samples had levels above the
quantitation limit of 10 ppt (0.085 |ig/m3). The highest reported concentration, monthly mean,
and mean composite for dibromochloromethane were 290 ppt (2.5 |ig/m3), 280 ppt (2.4 |ig/m3),
and 50 ppt (0.43 |ig/m3), respectively; all were recorded in downtown Los Angeles in June. Only
two monthly means were above 160 ppt; the remainder of the monthly means were below 60 ppt.
For bromoform, thirty-one percent of the samples had concentrations above the quantitation limit
of 10 ppt (0.10 |ig/m3). Peaks in the concentration of bromoform were observed at various sites
in May and June, with the downtown Los Angeles site registering the highest composite mean (40
ppt; 0.41 |ig/m3) and the highest monthly mean (310 ppt; 3.2 |ig/m3) in June 1983. Only two
monthly means were greater than 160 ppt; the remainder of the monthly means were below 60
ppt.
Atlas and Schauffler (1991) collected replicate air samples at various locations on the
Island of Hawaii during a month-long field experiment to test an analytical method for
determining halocarbons in ambient air. Dibromochloromethane was found at a mean level of
0.27 ppt, and bromoform was found at a mean concentration of 1.9 ppt. Information on sample
size and detection limit were not provided in the secondary source that reported this study (U.S.
EPA 1994b).
Wallace et al. (1982) conducted a pilot study designed to field test personal air-quality
monitoring methods. Personal air samples were collected from students at two universities:
Lamar University, Texas, located near a petrochemical manufacturing area, and the University of
North Carolina (UNC), located in a nonindustrialized area. The samples were analyzed for a
number of volatile organic compounds, including brominated trihalomethanes.
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Bromodichloromethane was detected in 64% of personal air samples from 11 Lamar students,
with a mean of 1.23 |ig/m3 (0.18 ppb), a median of 1 |ig/m3 (0.15 ppb), and a range of 0.12-3.72
|ig/m3 (0.018-0.56 ppb). The limit of detection was 0.24 |ig/m3 (0.036 ppb). At UNC, 17% of
the samples from 6 students had detectable levels of bromodichloromethane. Concentrations
ranged from 0.12-4.36 |ig/m3 (0.017-0.65 ppb) (mean, 0.83 |ig/m3 (0.12 ppb); median,
0.12 |ig/m3 (0.017 ppb)). Based on the above information, the average daily intake of
bromodichloromethane from air using an inhalation rate of 20 m3/day was estimated to be 25
|ig/day for Lamar students and 17 |ig/day for UNC students. Dibromochloromethane was not
present above 0.12 |ig/m3 (0.018 ppb) at either site.
b. Concentrations in Indoor Air
Relatively few studies have reported the concentrations of trihalomethanes in the indoor
air of homes. Kostiainen (1995) identified over 200 volatile organic compounds in indoor air of
26 houses identified by residents as causing symptoms such as headache, nausea, irritation of the
eyes, drowsiness, and fatigue. Bromoform was detected at low (unspecified) levels in 54 percent
of the homes, and no mention was made of dibromochloromethane or bromodichloromethane.
Weisel et al. (1999) measured brominated trihalomethane concentrations in indoor air in
New Jersey residences selected to examine low and high levels of drinking water contamination
with trihalomethanes. Descriptive statistics for trihalomethane concentration in water were
provided for the combined high and low concentration groups, but not for the individual
categories. One valid 15-minute air sample was collected at each of 48 residences. The indoor
air concentrations of bromodichloromethane averaged 0.38 ± 0.82 (SD) |ig/m3 (0.057 ± 0.12 ppb)
and 0.75 ± 0.96 |ig/m3 (0.11 ± 0.14 ppb) from the low and high water concentration groups,
respectively. The detection frequencies were 12/25 and 16 /23 in the low and high water
concentration groups, respectively. The indoor air concentrations of dibromochloromethane
averaged 0.44 ± 0.95 |ig/m3 (0.052 ±0.11 ppb) and 0.53 ± 0.84 |ig/m3 (0.062 ± 0.09 ppb) from
the low and high water concentration groups with detection frequencies of 5/25 and 7/23,
respectively. For bromoform, the average concentrations from the low and high water
concentration groups were 0.29 ± 0.93 |ig/m3 (0.028 ±0.089 ppb) and 0.35 ± 0.94 |ig/m3 (0.034 ±
0.091 ppb), with detection frequencies of 8/25 and 4/23, respectively. It was not clear whether
the averages were based on all measured samples or only those samples that were above the
detection limit for each compound.
Kerger et al (2000) evaluated the transfer of bromodichloromethane and
dibromochloromethane to indoor air in bathrooms during showering and bathing in homes
supplied with chlorinated tap water. The test sites were three urban homes containing three
bedrooms, a full bath, and approximately 1000 square feet of living space. The compounds were
simultaneously measured in hot and cold tap water (drawn from the kitchen sink) and in the
shower/bath enclosure and bathroom vanity area. Three shower protocols were examined: 6.8
min unventilated shower; 12 min unventilated shower and 6.8 min ventilated shower. Water flow
rate and temperature were monitored but not controlled. Airborne vapor samples were captured
by Summa canister and measured by gas chromatography using electron capture detection
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according to U.S. EPA method TO-14. Air samples were collected before, during and after the
water use event, for a total of 16 showers and 7 baths. Data for several events were eliminated
because of technical difficulties. For all shower protocols combined (n = 12), the increase in
average airborne concentration (± standard error), expressed as |ig/m3, in shower enclosure or
bathroom air per |ig/L in water, was 1.8 ± 0.3 for bromodichloromethane and 0.5 ±0.1 for
dibromochloromethane. For baths (n = 4), the average concentration increase during the bath was
0.59 ± 0.21 for bromodichloromethane and 0.15 ± 0.05 for dibromochloromethane. The relative
contribution of each chemical was consistent with the relative concentration in water and its
chemical and physical properties. The average exposures measured in this study were
approximately 30% lower than results reported by other investigators using EPA analytical
methods when data were normalized for water concentration, flow rate, shower volume, and
duration. This difference may have resulted from differences in the air exchange rate between
residential showers and laboratory test showers. These data are not adequate for characterizing
levels of individual brominated trihalomethanes in the home because the measurements targeted a
specific area of the residences and the sample size consisted of only three homes.
c. Estimates of Exposure from Air
The data available for occurrence of brominated trihalomethanes in air do not permit
calculation of a nationally aggregated intake estimate for the U.S. general population. To
accurately estimate total daily inhalation exposures, factors including location and season, the
fraction of time spent indoors compared with outdoors, potential exposures of individuals while
showering or bathing, potential exposure from volatilization of brominated trihalomethanes during
other household activities (e.g., use of dishwashers, toilet flushing), exposures of individuals who
spend large amounts of time at indoor pools, and potential for occupational exposures (e.g., for
laundromat or sewage treatment plant workers) require consideration. Although the existing data
do not permit such a refined analysis, they may be used to roughly estimate intake from air.
Based on the grand means calculated for multiple sampling locations by Brodzinsky and Singh
(1983), exposure to bromodichloromethane, dibromochloromethane and bromoform resulting
from inhalation of outdoor air can be roughly estimated assuming an inhalation rate of 20 m3/day,
100% absorption, and exposure to outdoor air for a full 24 hours per day. Using the mean
ambient air concentration of 110 ppt (0.74 |ig/m3) by volume for all sites reported in Brodzinsky
and Singh (1983), the daily intake of bromodichloromethane from outdoor air would be
21 |ig/day. Assuming a mean air concentration of 3.8 ppt (0.032 |ig/m3) for
dibromochloromethane, daily intake would be 0.64 |ig/day,. Assuming a mean air concentration
of 3.6 ppt (0.037 |ig/m3) by volume or bromoform, the daily intake would be 0.74 |ig/day.
Because these estimates are based on data from urban/suburban and industrial sites only, they may
represent high end exposures.
Adequate, nationally aggregated occurrence data are not available for calculating intake of
brominated trihalomethanes from indoor air. The indoor air concentrations measured by Weisel et
al. (1999) were not used for intake calculations because it could not be determined how the means
for each compound were calculated (i.e., whether all measurements were averaged or only those
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above the detection limit). In addition, the data were based on a single 15 minute air sample
collected from each of 48 homes located in a single state.
While brominated trihalomethane concentrations might be expected to be higher in indoor
air than in outdoor air due to confined space and additional indoor air sources (e.g. volatilization
from showering, baths, and other household activities), the available data do not allow such a
comparison.
Based on data from personal air monitors, Wallace et al. (1982) estimated daily inhalation
of bromodichloromethane to be 25 |ig/day for 11 students attending a university located near a
petrochemical manufacturing area and 17 |ig/day for 6 students attending a university in a
nonindustrialized area.. The personal air monitors registered bromodichloromethane from both
indoor (with the exception of showering and bathing) and outdoor exposures.
Dibromochloromethane was not detected and no data were available for bromoform.
3. Concentrations and Exposures Associated with Swimming Pools and Hot Tubs
Numerous studies have reported data for concentrations of brominated trihalomethanes
and exposures associated with swimming pools and hot tubs. Exposure of swimmers or hot tub
users to brominated trihalomethanes may result from dermal, ingestion, and inhalation exposure.
When evaluating these data, it is important to note that additional disinfectants are routinely
added to water contained in swimming pools and hot tubs; therefore, the levels of brominated
trihalomethanes present may not be representative of those in tap water.
Armstrong and Golden (1986) measured bromodichloromethane, dibromochloromethane,
and bromoform concentrations in the water and surrounding air of four indoor swimming pools,
five outdoor swimming pools, and four hot tubs. Concentrations in air were measured two
centimeters from the water surface. The bromodichloromethane concentrations of water in the
outdoor pools ranged from 1 to 72 |ig/L (ppb) (mean, 33 |ig/L). Levels in the indoor pools
ranged from 1 to 90 |ig/L (ppb) (mean, 16 |ig/L). The levels of bromodichloromethane in the hot
tubs ranged from <0.1 to 105 |ig/L (ppb) (mean, 17 |ig/L). Means and ranges of the bromo-
dichloromethane concentration two meters above the water surface for outdoor pools, indoor
pools, and hot tubs, respectively, were: <0.1 |ig/m3 (<0.015 ppb) (range not reported), 1.7 |ig/m3
(0.25 ppb) (range <0.1-10 |ig/m3 (0.015 -1.5 ppb)), and 1.4 |ig/m3 (0.21 ppb) (range <0.1-10
|ig/m3 (0.015 -1.5 ppb)). The dibromochloromethane concentration of water in the outdoor
pools ranged from <0.1 to 8 |ig/L (ppb) (mean, 4.2 //g/L (ppb)). Levels in the indoor pools
ranged from 0.3 to 30 |ig/L (ppb) (mean, 9.5 |ig/L (ppb)). The level of dibromochloromethane in
the hot tubs ranged from <0.1 to 48 |ig/L (ppb) (mean, 14.4 |ig/L (ppb)). Means and ranges of
the dibromochloromethane concentration two meters above the water surface for outdoor pools,
indoor pools, and hot tubs, respectively, were: <0.1 |ig/m3 (<0.01 ppb) (range not reported), 0.9
|ig/m3 (0.11 ppb) (<0.1-5 |ig/m3 (0.012-0.59 ppb)), and 0.7 |ig/m3 (0.08 ppb) (<0.1- 5 |ig/m3
(0.012-0.59 ppb)). The mean bromoform concentration in the outdoor pools was less than
0.1 |ig/L (ppb). Levels in the indoor pools ranged from less than 0.1 to 20 |ig/L (ppb) (mean,
6 |ig/L (ppb)). The levels of bromoform in the hot tubs ranged from less than 0.1 to 62 |ig/L
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(ppb) (mean, 13 |ig/L (ppb)). Means and ranges of the bromoform concentration two meters
above the water surface for outdoor pools, indoor pools, and hot tubs, respectively, were: <0.1
|ig/m3 (<9.7 ppt) ( range not reported), 9 |ig/m3 (870 ppt) (<0.1-14 |ig/m3 (9.7-1360 ppt)), and
8 |ig/m3 ( 770 ppt) (<0.1-14 |ig/m3 (9.7-1360 ppt)).
Cammann and Hiibner (1995) compared concentrations of trihalomethanes in swimmers'
and bath attendants' blood and urine before and after swimming or working in indoor swimming
pools. Water and air concentrations were measure in different locations in the pool environment.
The purpose was to determine whether blood levels of trihalomethanes would reflect inhalation
exposure to trihalomethanes in the pool environment and whether those compounds also would
appear in urine. Measured concentrations of bromodichloromethane, dibromochloromethane, and
bromoform in samples of swimming pool waters collected at a depth of 10 to 20 cm were 0.69 to
5.64 |ig/L (ppb), 0.03 to 6.51 |ig/L (ppb), and 0.14 to 2.32 |ig/L (ppb) for the three compounds,
respectively. Averages (± SD) of the 10 pool water measurements presented in Table 1 of the
report were 2.12 ± 1.52 |ig/L (ppb), 1.11 ± 2.07 |ig/L (ppb), and 0.42 ± 0.73 |ig/L (ppb) for
bromodichloromethane, dibromochloromethane, and bromoform, respectively. Average (± 1 SD)
concentrations in the four air samples taken (location of sampling not specified) were 15.4 ± 7.36
|ig/m3 (2.30 ±1.10 ppb), 1.94 ± 1.01 |ig/m3 (0.228 ±0.119 ppb), and below the quantitation limit
(QL) (not specified, although probably 0.02 ppb) for bromodichloromethane,
dibromochloromethane, and bromoform, respectively.
Measurements of bromodichloromethane in 8 bath attendants' blood before their shifts
ranged from below QL for 12/18 measurements (67%) to 0.1 |ig/L (ppb) (Camman and Hiibner,
1995). After their shifts, the concentrations ranged from below QL in 7/18 measurements (39%)
to 0.6 jUg/L (ppb). Similarly, measurements of bromodichloromethane in swimmers' blood was
higher after than before swimming. Before swimming, blood concentrations of
bromodichloromethane ranged from less than the QL in 10/20 (50%) swimmers to 0.2 |ig/L
(ppb); while after swimming, blood concentrations were above the QL in all 20 swimmers,
ranging from « 0.02 to 0.4 /ig/L (ppb) in 19 of the swimmers. The twentieth swimmer had a
blood concentration of « 1.5 //g/L (ppb). For all but two of the swimmers, blood concentrations
of bromodichloromethane had dropped below the QL by the next day (values for the other two
swimmers were less than 0.1 //g/L (ppb)). Dibromochloromethane and bromoform were not
detected in the blood of either the bath attendants or swimmers. None of the brominated
trihalomethanes were detected in the urine of the study subjects. Thus, only exposure to
bromodichloromethane by inhalation (bath attendants) or inhalation, dermal absorption, and
ingestion (swimmers) is reflected in increased blood levels of the compound. Blood levels of
bromodichloromethane usually returned to pre-exposure levels within 24 hours after the exposure
Aggazzotti et al. (1998) evaluated concentrations of trihalomethanes in the blood and
breath of five competitive swimmers regularly training in an indoor swimming pool in Italy. The
group included three males and two females between the ages of 17 and 21 years. All were non-
smokers. Concurrent sampling of blood, alveolar air, and environmental air occurred at five times
for each of four sessions: (a) at the University Department two hours before arriving at the pool,
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(b) after one hour sitting near the edge of the pool, (c), after one hour of swimming, (d) back at
the University one hour after swimming ended, and (e) at the University 1.5 hr after swimming
ended. While bromodichloromethane and dibromochloromethane were always found in water and
environmental air samples at the pool immediately before and after the 1-hr swimming session,
bromoform was rarely detected in the indoor pool air. None of the three brominated
trihalomethanes were detected in the air at the University Department or in the alveolar air of the
swimmers at the Department two hours before arriving at the pool. At the pool, prior to the
swimming session, the means (± SD) of the four measured ambient air concentrations of
bromodichloromethane, dibromochloromethane, and bromoform werel0.5 ±3.1 |ig/m3 (1.6 ±
0.46 ppb; 4 detects), 5.2 ±1.5 |ig/m3 (0.61 ± 0.17 ppb; 4 detects), and 1 detect of 0.2 |ig/m3 (0.02
ppb), respectively. At the pool, just after the 1-hr swimming session, the means (± SD) of the
four measured ambient air concentrations of bromodichloromethane, dibromochloromethane, and
bromoform were 20.0 ±4.1 |ig/m3 (2.99 ±4.1 ppb; 4 detects), 11.4 ±2.1 |ig/m3 (1.34 ± 0.23 ppb;
4 detects), and 1 detect of 0.2 |ig/m3 (0.02 ppb), respectively.
Concentrations of bromodichloromethane and dibromochloromethane in the alveolar air of
the swimmers before and after the swimming session indicated inhalation uptake of both
compounds (Aggazzotti et al., 1998). At the pool, prior to the swimming session, the means (±
SD) of the 20 measured alveolar air concentrations (5 swimmers assessed at each of 4 sessions) of
bromodichloromethane and dibromochloromethane were 2.7 ±1.2 |ig/m3 (0.40 ±0.18 ppb) and
0.8 ± 0.8 |ig/m3 (0.09 ± 0.09 ppb), respectively. Bromoform was not detected in any of the 20
samples. At the pool, after the 1-hr swimming session, the means (± SD) of the alveolar air
concentrations of bromodichloromethane and dibromochloromethane were 6.5 ±1.3 |ig/m3 (0.97
±0.19 ppb) and 1.4 ± 0.9 |ig/m3 (0.16 ± 0.11 ppb), respectively. Bromoform was not detected in
any of the 20 samples. Blood levels of bromodichloromethane and dibromochloromethane before
and after swimming, on the other hand, were below detection limits in most samples, and hence
showed no trends.
Aggazzotti et al. (1998) estimated uptake of the trihalomethanes of the resting and active
swimmers using the following assumptions. At rest, the pulmonary ventilation rate of the women
was 6 liters per minute (L/min) while that of men was 7.5 L/min. During swimming, the
ventilation rate of the women was 25 L/min while that of the men 36 L/min. The estimated
uptake rates of bromodichloromethane for the five swimmers at rest ranged from 2.8 to 3.7 //g
per hour (/xg/h), with a mean value of 3.3 ± 0.41(SD) //g/h for the three males and two females
combined. The estimated uptake rates for the same individuals actively swimming were 20 to 30
//g/h, with a mean value of 26 ± 5.1 //g/h. The estimated uptake rates of dibromochloromethane
for the five swimmers at rest ranged from 1.5 to 2.0 //g/h, with a mean value of 1.8 ± 0.23 //g/h.
The estimated uptake rates during swimming increased to between 14 and 22 //g/h, with a mean
value of 18 ± 3.6 //g/h. Occurrence of dermal uptake was acknowledged but not estimated.
Lindstrom et al. (1997) also assessed exposure of two competitive swimmers to
bromodichloromethane during training sessions at an indoor pool. The indoor pool air
concentrations of bromodichloromethane collected over 60- and 119-minute intervals were 2.76
and 3.02 |ig/m3 (0.41 and 0.45 ppb), respectively. Breath samples were collected from the
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swimmers before, during, and for 3 hours after a training workout. Breath samples collected
during the workout demonstrated a rapid uptake of bromodichloromethane to maximum alveolar
concentrations of 5 to 6 |ig/m3 (0.7 to 0.9 ppb), which are higher than the ambient air
concentrations. The authors concluded that significant (80% of total exposure) dermal absorption
of the related trihalomethane chloroform from water was occurring, but did not estimate the
extent of dermal uptake for bromodichloromethane.
4. Soil Concentrations and Exposure
Data on the concentration of brominated trihalomethanes in soil were not available in the
materials reviewed for this document. Based on the measured Henry's Law constant and vapor
pressure of the individual compounds, volatilization from both wet and dry soil surfaces should be
relatively rapid (U.S. EPA 1987). Therefore, exposure from soil ingestion is not considered to be
a significant route for exposure to the brominated trihalomethanes.
C. Overall Exposure
The RSC (relative source contribution) is the percentage of total daily exposure that is
attributable to tap water when all potential sources are considered (e.g., air, food, soil, and
water). Ideally, the RSC is determined quantitatively using nationwide, central tendency and/or
high-end estimates of exposure from each relevant medium. In the absence of such data, a default
RSC ranging from 20% to 80% may be used.
The RSC used in the current and previous drinking water regulations for
dibromochloromethane is 80%. This value was established by use of a screening level approach
to estimate and compare exposure to dibromochloromethane from various sources. Information
considered for during this process is summarized in Appendix C. The use of the 80% value for
the RSC for dibromochloromethane is supported by limited use of this chemical in industrial
applications with potential for direct release to the environment. The use of the 80% value is
further supported by apparently low concentrations in foods and soils and the potential for human
exposure to dibromochloromethane in tap water via three exposure routes: 1) ingestion as
drinking water; 2) inhalation of volatilized dibromochloromethane during use of tap water for
household activities; and 3) by dermal exposure during showering, bathing, or other activities.
The available data for concentrations of outdoor air and food, although limited, suggest that
exposures via these routes are likely to be low when compared to water.
Parallel RSC calculations were not performed for bromodichloromethane and bromoform.
The EPA has set the regulatory level for these chemicals in drinking water at zero because it has
been determined that they are probable human carcinogens. Therefore, determination of an RSC
is not relevant for these chemicals because it is the Agency's policy to perform RSC analysis only
for noncarcinogens.
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D. Body Burden
1. Blood and Breath Levels
Barkley et al. (1980) analyzed blood samples from nine residents of the old Love Canal
area in 1978 for a variety of volatile organic compounds, including all three of the brominated
trihalomethanes. Bromodichloromethane was detected in the blood of one individual; its
concentration was 14 |ig/L (ppb). Dibromochloromethane and bromoform were not detected.
Antoine et al. (1986) analyzed the blood of 250 environmentally sensitive patients for 18
volatile organic compounds. Bromoform concentrations ranged from undetectable to 3.4 |ig/L
(ppb), with a mean of 0.6 |ig/L (ppb).
Ashley et al. (1994) analyzed samples of whole blood of 600 or more people in the United
States who participated in the Third National Health and Nutrition Examination Survey
(NHANES III) for 32 volatile organic compounds using analytical methods designed to measure
extremely low concentrations. Bromodichloromethane, with a detection limit of 0.009 |ig/L (ppb)
was detected only in 14% of 1072 samples. Dibromochloromethane, with a detection limit of
0.013 |ig/L (ppb), was detected in only 12% of 1035 samples. Using unprocessed commercial
Vacutainer Tubes, Ashely et al. (1994) initially obtained measures of bromoform concentrations in
blood similar to those reported by Antoine et al. (1986). However, using Vacutainer Tubes that
had been processed to removed VOCs prior to use, Ashely et al. (1994) detected bromoform in
less than 10% of samples analyzed at a detection limit of 0.027 |ig/L (ppb). Wallace (1997)
obtained the summary statistics for bromodichloromethane and dibromochloromethane, which
were not published in Ashely et al.'s (1994) paper. The mean (± SD) of the measured blood
concentrations were 0.0077 ± 0.0178 and 0.00886 ± 0.00856 |ig/L (ppb), respectively. The
median values were below the limit of detection. The upper 90th percentile values were 0.0122
and 0.0151 |ig/L (ppb), respectively.
Weisel et al. (1999) measured brominated trihalomethanes in the exhaled breath of female
subjects after showering. The study authors recruited 49 women who had previously participated
in a case-control study on neural tube birth defects from locations throughout the state of New
Jersey (Klotz and Pyrch, 1999). The method used to select the subjects provided a wide range of
brominated trihalomethane exposures within the home, in contrast to a distribution of exposures
that might exist within a single water distribution system or within the general population.
Exposure to brominated trihalomethanes was estimated by collection of duplicate cold tap water
samples, collection of a 15-minute air sample, and responses to a 48-hour recall questionnaire on
water use in the home. Post-shower whole breath samples were collected by having the subject
blow into a Tedlar® sampling bag at the conclusion of a shower. Background breath samples
were collected at a subsequent home visit by the investigators. Valid samples were obtained from
33 of the subjects. However, the time of post-shower sample collection as reported by the
subjects varied from immediately after the shower to 20 minutes later. As noted by the authors,
the delay in sample collection is an important determinant in breath concentrations because
trihalomethane breath concentration declines exponentially after exposure ceases. As a result,
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each subject was assigned to one of three groups: 1) breath sample collected within 5 minutes
after completion of shower (Group A; n = 13); 2) breath sample collected within 5 to 20 minutes
after completion of shower (Group B; n=14); or 3) breath sample collected more than 20 minutes
after showering (Group C; n=6). The breath concentrations on individual brominated
trihalomethanes for each group were compared to measured water concentrations and estimates
of exposure (calculated as the product of the water concentration and reported duration of the
shower; shower duration data and calculated exposure estimates were not reported).
The mean (± standard deviation) concentrations of bromodichloromethane,
dibromochloromethane, and bromoform were 5.7 ± 8.6, 2.0 ± 2.1, and 0.73 ± 0.90 //g/L (ppb),
respectively. The median values for the three compounds were 2.6, 1.4, and 0.45 //g/L (ppb),
respectively. Bromodichloromethane showed significant correlations for breath and water
concentration and breath and shower exposure for Groups A and B. Significant correlations for
dibromochloromethane and bromoform were found for Group A participants. Analytical
variability related to low concentrations of dibromochloromethane and bromoform (near the
detection limit) may have obscured trends in the data for Group B. source of in the houses and
found significant correlations between the water concentration of each brominated trihalomethane
and the concentration of that trihalomethane in expired air if the air samples were collected within
5 minutes of showering. Results of statistical analysis for Group C were not reported because the
sample size was small and the authors considered the results questionable. The observed results
were considered consistent with showering being a source of exposure to brominated
trihalomethanes.
Backer et al. (2000) examined levels of brominated trihalomethanes in whole blood
following three types of water use events by adult volunteers: showering for 10 minutes in tap
water (n=l 1); bathing for 10 minutes in a tub filled with tap water (n=10); or consumption of one
liter of tap water over a 10 minute period (n=10). Each participant provided a blood sample
immediately before exposure, 10 minutes after exposure ended, and 30 minutes (showering and
bathing) or one hour (ingestion) after exposure. Tap water and blood samples were analyzed by
purge-and-trap/gas chromatography/mass spectrometry with detection capability in the parts per
quadrillion range. Bromoform was not detected in either tap water or whole blood. Mean tap
water concentrations of bromodichloromethane and dibromochloromethane were 6 |ig/L and 1.1
|ig/L, respectively. The highest levels of these compounds in whole blood occurred 10 minutes
after exposure had ended. The second post-exposure measurements showed that blood levels of
both compounds had decreased, but were still above the pre-exposure baseline levels in subjects
who took showers or baths. Measurement data are shown in Table IV-9 below.
The study authors reported that similar relative findings were obtained for dibromochloromethane
(data shown graphically in the study report). These data indicate a dramatic difference between
the whole blood levels resulting from ingestion and those resulting from bathing or showering
(including dermal, inhalation, and possibly ingestion exposure). Blood level increases observed
for each compound after ingestion of one liter of water were less than 10% of those observed
after bathing or showering for 10 minutes. The blood level increases observed for each
compound 10 minutes after bathing or showering were significantly
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Table IV-9 Mean Bromodichloromethane Concentrations in Blood Following Three Types
of Water Use Events
Water Use Event
Median Bromodichloromethane Concentration in Whole Blood (pg/mL)
Pre-exposure
10 minutes post-exposure
30 or 60 minutes post-
exposure
10 Minute Shower
3.3
19.4
10.3 (30 min)
10 Minute Bath
2.3
17.0
9.9 (30 min)
Ingestion of 1 L
2.6
3.8
2.8 (60 min)
increased (p<0.01) when compared to the post-ingestion blood levels. Measurable levels of
bromodichloromethane and dibromochloromethane in the pre-exposure whole blood samples
were attributed to recent prior exposure or to bioaccumulation after repeated exposure to tap
water.
In addition to the differences observed in whole blood levels of bromodichloromethane
and dibromochloromethane among exposure groups, the study authors observed that the blood
concentration data for each chemical occurred in two clusters within each exposure group. The
mean increases for the two clusters observed after bathing or showering were significantly
different for bromodichloromethane. The same individuals who had greater increases of
bromodichloromethane also experienced greater increases of dibromochloromethane and
chloroform in the blood after bathing or showering. The underlying basis for the observed
clustering is unknown, but was not related to gender. The study authors suggested that
polymorphic expression of a metabolizing enzyme (e.g. glutathione-S-transferase theta) or
differences in fitness level (resulting in inhalation of larger volumes of air) may have accounted for
the observed pattern. However, they noted that differences in fitness level would more likely be
expected to result in a continuous distribution.
Lynberg et al. (2001) conducted a field study in Corpus Christi, Texas, and Cobb County,
Georgia, to evaluate exposure measures for disinfection by-products, including brominated
trihalomethanes. These areas were selected for study based on the following criteria: 1) relatively
high trihalomethane concentrations relative to national averages; 2) high intrasystem differences
that would result in a potential exposure gradient across the study population; 3) one water
distribution system with predominately chlorinated species of trihalomethanes (i.e., chloroform)
and one water system with predominately brominated trihalomethanes; and 4) a water utility
service population large enough to allow rapid selection of 25 mothers per geographic area who
had given birth to healthy babies from June, 1998 through May, 1999. Exposure to individual
trihalomethanes was assessed by collection of blood and water samples and by collection of
information on water use patterns and tap water characteristics. Whole blood samples were
collected before and after showering. Levels of individual trihalomethanes were determined for
samples collected in the home of participants, in the distribution system, and at the water
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treatment plants. A modified version of the Total Exposure Model (TEM) was used to estimate
uptake of trihalomethanes into the bloodstream (data for chloroform exposure were presented for
one individual in Corpus Christi). The results of the study indicate that concentration of
individual trihalomethanes varied by site and location within the water system (Table IV-10). In
Corpus Christi water samples, brominated trihalomethanes accounted for 71% of the total
trihalomethane concentration by weight. In contrast, brominated trihalomethanes accounted for
only 12% of the trihalomethanes in Cobb County water samples. Significant differences (p =
0.0001) in the blood levels of dibromochloromethane and bromoform were observed between
study locations (Table IV-11). The differences between locations were evident both before and
after showering. The study authors indicated that there was considerable variability in blood
levels of trihalomethanes among participants from a single location. For example, pre-shower
chloroform blood levels in Cobb County ranged from 130 ppt to 1100 ppt. No data were
presented for the brominated trihalomethanes. The variability was tentatively attributed to
different patterns of household water use among participants. Significant increases (p = 0.0001)
in blood levels of all brominated trihalomethanes were observed after showering. The increases in
dibromochloromethane and bromoform were significantly greater in Corpus Christi than in Cobb
County. No TEM modeling data were presented for brominated trihalomethanes. However,
TEM results presented for chloroform exposure for one study participant who consumed bottled
water indicated that inhalation exposure in the household accounted for approximately 98% of the
calculated 24-hour chloroform dose, with the remainder attributed to the dermal route. Overall,
this study demonstrates that blood levels of brominated trihalomethanes vary significantly across
populations, with water quality characteristics and water use activities being important variables.
Table IV-10 Median Tap Water Trihalomethane Levels (ppb) in Cobb County and Corpus
Christi Homes, Water Treatment Plants, and Distribution Systems
Trihalomethane
Cobb County
Corpus Christi
Home
(n=25)
Distribution
System
(n=20)
Water
Treatment
Plant
(n=7)
Home
(n=25)
Distribution
System
(n=30)
Water
Treatment
Plant
(n=20)
Bromodichloromethane
13.5
12.5
9.5
12.2
8.3
9.5
Dibromochloromethane
1.7
2.4
1.4
13.5
12.6
14.3
Bromoform
NDa
ND
ND
8.7
9.7
11.9
Chloroform
84.8
79
49.5
8.2
4.6
6.7
¦' ND, not detected (detection limit < lppb)
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Table IV-11 Between Site Comparison of Median Blood Levels (ppt) and Changes in
Blood Levels (ppt) after Showering
Trihalomethane
Before Shower
After Shower
Change in Blood Level after
Showering
Cobba
Corpus13
Cobb
Corpus
Cobb
Corpus
Bromodichloromethane
6.2
6.8
38
43
30
34
Dibromochloromethane
1.2
7.0
6.1
41
5.0
35
Bromoform
0.3
3.5
0.5
17
0.2
12
Chloroform
70
25
280
57
189
25
a Cobb County, Georgia
b Corpus Christi, TX
2. Mother's Milk
Pellizzari et al. (1982) analyzed the milk of eight nursing mothers for various compounds,
including bromodichloromethane and dibromochloromethane. The samples were collected from
49 lactating women living in the vicinity of chemical manufacturing plants and/or industrial user
facilities in Bridgeville, PA, Bayonne, NJ, Jersey City, NJ, and Baton Rouge LA. Both
compounds were identified in one of the eight samples. Actual concentrations and detection
limits were not reported. Kroneld and Reunanen (1990) did not detect any of the brominated
trihalomethanes in human milk in a study conducted in Turku, Finland.
E. Summary
Brominated trihalomethanes are found in virtually all water treated for drinking; however,
concentrations of individual forms vary widely depending on the type of water treatment, locale,
time of year, sampling point in the distribution system, and source of the drinking water.
Occurrence data for brominated trihalomethanes are available from 13 national surveys and 9
additional studies that are more restricted in scope. The procedures used for sampling processing
and storage and calculation of summary statistics should be carefully considered when evaluating
and comparing brominated trihalomethane occurrence data. Some methods restrict
trihalomethane formation by refrigeration or the use of quenching agents, whereas others
maximize trihalomethane formation by storage at room temperature. Approaches to data
summarization vary by study in the treatment of data below the analytical detection level or
minimum reporting level.
When all available national survey data are considered, bromodichloromethane concen-
trations in drinking water range from below the detection limit to 183 |ig/L (ppb), while dibromo-
chloromethane and bromoform concentrations range from below the detection limit to 280 |ig/L
(ppb). When data for the three brominated trihalomethanes are compared, the frequency of
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detection and measured concentrations of bromodichloromethane in drinking water supplies tend
to be higher than those for dibromochloromethane. Bromoform is detected less frequently and at
lower concentrations than the other two brominated trihalomethanes, except in some ground
waters. Concentrations of all trihalomethanes in drinking water were generally lower when the
raw water is obtained from ground water sources rather than surface water sources. The most
recent national survey data are those collected by the U.S. EPA under the Information Collection
Rule (ICR). Monitoring data were collected over an 18-month period between July 1997 and
December 1998 from approximately 300 water systems operating 501 plants and serving at least
100,000 people. Summary occurrence data stratified by raw water source (groundwater or
surface water) are available for finished water, the distribution system (DS) average, and the DS
high values. The mean, median, and 90th percentile values for surface water DS average
concentrations in the ICR survey are 8.6, 70.2, and 20.3 |ig/L, respectively, for
bromodichloromethane (range of individual values 0 - 65.8 |ig/L); 2.4, 4.72, and 13.2 |ig/L,
respectively, for dibromochloromethane (range 0 - 67.3); and 0. 1.18, and 3.10, respectively, for
bromoform (range 0-3.43).
Relatively few studies have analyzed non-beverage foods for the occurrence of brominated
trihalomethanes. In the few studies available, bromodichloromethane has been detected in non-
beverage foods (i.e., in one sample of butter at 7 ppb, in three samples of ice-cream at 0.6 to 2.3
ppb, in 6 of 10 samples of bean curd at 1.2 to 5.2 ppb, and in one sample of bacon (probably
below the minimal quantitation limit)). In addition, bromodichloromethane was detected in one
sample each of eleven foods out of 70 tested in 14 Market Baskets for the FDA Total Diet Study.
The detected concentrations ranged from 10 to 37 ppb for individual food items. Studies that
analyzed non-beverage foods for dibromochloromethane and bromoform detected neither
compound in any of the samples. Brominated trihalomethanes have been detected in up to a third
or one half of the types of prepared beverages examined in some studies, being detected most
frequently in colas and other carbonated soft drinks. Bromodichloromethane has been found most
frequently of the three compounds and bromoform the least frequently. Bromodichloromethane
was detected in approximately half of the prepared beverages examined by McNeal et al. (1995)
in the United States and in all of 13 soft drinks that they analyzed. With the exception of one of
the 13 soft drinks examined by McNeal et al. (1995) with a concentration of 12 ppb, none of the
at least 18 other measured concentrations of bromodichloromethane in soft drinks described
above (three from Entz et al. (1982), three from Uhler and Diachenko (1987), and the remaining
12 from McNeal et al. (1995)) exceeded a value of 4 ppb. Bromodichloromethane was detected
in one sample of fruit juice at 5 ppb.
Exposure to brominated trihalomethanes via ingestion of drinking water was estimated
using data obtained for disinfectants and disinfection byproducts under the Information Collection
Rule (ICR). ICR data offer several advantages over other national studies for purposes of
estimating national exposure levels of adults in the United States to brominated trihalomethanes
via ingestion of drinking water. First, they are recent and reflect relatively current conditions.
Second, data of very similar quality and quantity were collected systematically from a large
number of plants (501) and systems (approximately 300), including both surface and ground
water systems. Third, the mean, median, and 90th percentile value were estimated on the basis of
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all samples taken, not just the sample detects. Thus, these descriptive statistics are representative
of the exposures of the entire populations served by those systems, not just the populations served
by systems with higher concentrations of these compounds. However, this study can not be
considered representative of smaller public water supplies or water supplies from the most highly
industrialized or contaminated areas.
Exposure was calculated by multiplying the concentration of individual brominated
trihalomethanes in drinking water by the average daily intake, assuming that each individual
consumes two liters of water per day. The annual median, mean, and upper 90th percentile values
are presented for both surface and ground water systems. Assuming that the DS High value
actually represents the average exposure level of persons served by one plant distribution pipe
with the longest water-residence time, the DS High value might be used to estimate a high-end
exposure level.
For bromodichloromethane, the median, mean, and 90th percentile population exposures
from surface water systems are estimated to be 17, 20, and 40 |ig/person/day, respectively. The
same values for populations exposed to bromodichloromethane from ground water systems are
lower - 3.6, 8.1, and 22 |ig/person/day, respectively. For dibromochloromethane, the median,
mean, and 90th percentile population exposures from surface water systems are estimated to be
4.8, 9.4, and 26 |ig/person/day, respectively. The corresponding values for populations exposed
to dibromochloromethane from groundwater system are lower - 2.7, 6.2, and 18 |ig/person/day,
respectively. For bromoform, the median, mean, and 90th percentile population exposures from
surface water systems are estimated to be near 0, 2.4, and 6.2 |ig/person/day, respectively. The
same values for populations exposed to bromoform from ground water systems are higher - 0.65,
3.8, and 9.6 |ig/person/day, respectively.
For purposes of comparison, estimates of ingestion exposure to bromodichloromethane,
dibromochloromethane, and bromoform in drinking water were also estimated from data collected
in other, older studies. Ingestion from ground water supplies was estimated from the median
levels found in the Ground Water Supply Survey conducted by U.S. EPA in 1980-81. Based on
the range of median levels (1.4-2.1 |ig/L (ppb)) and a consumption rate of two liters per day, the
median ingestion exposure to bromodichloromethane may range from 2.8 to 4.2 |ig/day.
Similarly, median exposure to dibromochloromethane may range from 4.2 to 7.8 |ig/day, and for
bromoform, median exposure may range from 4.8 to 8.4 |ig/day. Exposure to
bromodichloromethane from surface water supplies can be estimated based on the range of
median values observed under different conditions in the National Organics Monitoring Survey
conducted by U.S. EPA in 1976-1977, which mainly sampled surface water systems. Based on a
range of 5.9-14 |ig/L (ppb), exposure to bromodichloromethane from surface water is estimated
to be between 12 and 28 |ig/day. Similarly, based on the range of medians reported for
dibromochloromethane concentrations, the median exposure is estimated to be up to 6 |ig/day.
The median levels of bromoform in the surface water supplies have been found to be less than the
EPA Drinking Water minimum reporting levels (MRLs) of 0.5-1 |ig/L (ppb). An estimate of
exposure based on the MRLs will be overly conservative because the actual concentration of
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bromoform is not detectable. Based on the range of MRLs, 0.5-1 |ig/L (ppb), the exposure to
bromoform is estimated to range from 1 to 2 |ig/day for surface water supplies.
Ingestion exposure to brominated trihalomethanes in drinking water can also be estimated
from the concentrations found at the tap in the U.S. EPA's Total Exposure Assessment
Methodology (TEAM) study. Estimates of the average of the population intakes for ingestion of
bromodichloromethane from drinking water range from 0.42 to 42 /ig/person/day. The upper 90th
percentile estimates range from <2.0 to 90 //g/person/day. Estimates of the average population
intake of dibromochloromethane from drinking water range from 0.2 to 56 /ig/person/day. The
upper 90th percentile estimates range from < 0.9 to 86 //g/person/day. Estimates of the average of
the population intakes of bromoform, for those areas in which bromoform was measurable in a
majority of the samples, range from 1.6 to 16.2 //g/person/day. The upper 90th percentile
estimates range from 2.4 to 26 /ig/person/day. Four of the six locations in the TEAM study,
however, had a low frequency (less than 10%) of detection of bromoform in measurable
quantities.
Sources of uncertainty in these estimates of ingestion exposure include use of different
analytical methods, failure to report quantitation limits, using measures near the detection limit,
failure to report how nondetects are handled when averaging values (e.g., set to zero or one half
the detection limit), and failure to report sample storage method and duration. In addition, many
environmental factors influence the concentrations of these compounds in drinking water at the
tap and in vended or bottled waters used for drinking. These factors include season and
temperature, geographic location, source of water, residence time in distribution system, and
others.
Average daily intake of dibromochloromethane via ingestion, dermal contact, and
inhalation of compound volatilized during household use were also estimated for determination of
the Relative Source Concentration (RSC). Intake for ingestion was calculated using mean intake
rates of 1.2 or 0.6 L/day for total and direct intake (NRC, 1999), respectively. Direct intake
includes consumption of water directly from the tap, but does not include intake of tap water used
for preparation of heated items such tea, coffee, or soup. Based on the ICR distribution system
average concentration of 4.72 |ig/L for dibromochloromethane in surface water, the average daily
total and direct and ingestion intakes would be 5.7 and 2.8 |ig/day, respectively. The average
dermal uptake of dibromochloromethane was estimated to be 2 |ig per shower or bathing event.
Average daily intake via inhalation of dibromochloromethane volatilized during to be 7 |ig/day for
the volatilized compound. Parallel calculations were not performed for bromodichloromethane or
bromoform, because these compounds are probable carcinogens. Therefore, in accordance with
U.S. EPA policy, RSC analysis was not conducted.
Some data on the occurrence of brominated trihalomethanes in foods and beverages are
available from studies conducted in Italy, Japan, and Finland. These studies were also limited in
scope to examination of relatively few food or beverage items. Bromodichloromethane,
dibromochloromethane, and bromoform concentrations measured in foods and beverages in Italy,
Japan and Finland ranged from undetectable to 40 ppb, undetectable to 13.9 ppb, and
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undetectable to 10.7 ppb, respectively. Because of possible differences in water disinfection or
food processing practices, these data may not be representative of concentrations in foods and
beverages produced in the U.S.
Concentrations in outdoor air were variable from site to site. When data from several
urban/suburban and source-dominated sites in Texas, Louisiana, North Carolina and/or Arkansas
were combined, the resulting average outdoor air concentrations were 110 ppt (0.74 |ig/m3) for
bromodichloromethane, 3.8 ppt (0.032 |ig/m3) for dibromochloromethane, and 3.6 ppt (0.037
|ig/m3) for bromoform. A regional study conducted at several sites in southern California found
bromodichloromethane, dibromochloromethane, and bromoform in 35%, 17%, and 31% of the
samples, respectively. The maximum concentrations observed were 40 ppt (0.27 |ig/m3) for
bromodichloromethane; 290 ppt (2.5 |ig/m3) for dibromochloromethane; 310 ppt (3.2 |ig/m3) for
bromoform. Bromodichloromethane was detected in 64% (n=l 1) and 17% (n=6) of personal air
samples collected in Texas and North Carolina. The detected concentrations ranged from 0.12 to
4.36 |ig/m3 (0.017 to 0.65 ppb). Dibromochloromethane was not detected.
Mean concentrations in indoor air ranged from 0.38 to 0.75 |ig/m3 for bromodichloro-
methane; 0.44 to 0.53 |ig/m3 for dibromochloromethane, and 0.29 to 0.35 |ig/m3 for bromoform,
as determined from 15 minute samples collected in 48 New Jersey residences. In a separate study,
levels of brominated trihalomethanes in indoor air were locally increased (e.g., in shower/bath
enclosures and vanity areas) during showering and bathing events. The levels of individual
brominated trihalomethanes in air were reported to be consistent with the levels in tap water.
The use of chlorine to disinfect swimming pools and hot tubs results in the formation of
brominated trihalomethanes. Swimming pool and hot tub users are potentially exposed to
brominated trihalomethanes via dermal contact, ingestion, and inhalation of compounds released
to the overlying air. As a result, swimming pool and hot tub users may experience greater overall
exposures to brominated trihalomethanes than the general population. One study indicated that
bromodichloromethane, dibromochloromethane, and bromoform concentrations in swimming pool
and hot tub water ranged from 1 to 105 |ig/L (ppb), from 0.1 to 48 |ig/L (ppb), and from less
than 0.1 to 62 |ig/L (ppb), respectively. Concentrations of the same brominated trihalomethanes
in the air two meters above the pool water ranged from less than 0.1 to 14 |ig/m3 (0.015-2.09
ppb), from less than 0.1-10 |ig/m3 (0.011-1.2 ppb), and from less than 0.1 to 5.0 |ig/m3
(0.0097-0.48 ppb), respectively. Data from several studies confirm the uptake of brominated
trihalomethanes from swimming pools and environs by dermal and/or inhalation pathways.
No data for occurrence of brominated trihalomethanes in soil were available in the
materials reviewed for this document. The chemical and physical properties of the brominated
trihalomethanes indicate that they should volatilize readily from wet or dry soil surfaces.
Therefore, ingestion of soil is not expected to be a significant route of exposure.
Exposure to brominated trihalomethanes via ingestion of drinking water was estimated
using data obtained for disinfectants and disinfection byproducts under the Information Collection
Rule (ICR). ICR data offer several advantages over other national studies for purposes of
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estimating national exposure levels of adults in the United States to brominated trihalomethanes
via ingestion of drinking water. First, they are recent and reflect relatively current conditions.
Second, data of very similar quality and quantity were collected systematically from a large
number of plants (501) and systems (approximately 300), including both surface and ground
water systems. Third, the mean, median, and 90th percentile value were estimated on the basis of
all samples taken, not just the sample detects. Thus, these descriptive statistics are representative
of the exposures of the entire populations served by those systems, not just the populations served
by systems with higher concentrations of these compounds. However, this study can not be
considered representative of smaller public water supplies or water supplies from the most highly
industrialized or contaminated areas.
Exposure was calculated by multiplying the concentration of individual brominated
trihalomethanes in drinking water by the average daily intake, assuming that each individual
consumes two liters of water per day. The annual median, mean, and upper 90th percentile values
are presented for both surface and ground water systems. Assuming that the DS High value
actually represents the average exposure level of persons served by one plant distribution pipe
with the longest water-residence time, the DS High value might be used to estimate a high-end
exposure level.
For bromodichloromethane, the median, mean, and 90th percentile population exposures
from surface water systems are estimated to be 17, 20, and 40 |ig/person/day, respectively. The
same values for populations exposed to bromodichloromethane from ground water systems are
lower - 3.6, 8.1, and 22 |ig/person/day, respectively. For dibromochloromethane, the median,
mean, and 90th percentile population exposures from surface water systems are estimated to be
4.8, 9.4, and 26 |ig/person/day, respectively. The corresponding values for populations exposed
to dibromochloromethane from groundwater system are lower - 2.7, 6.2, and 18 |ig/person/day,
respectively. For bromoform, the median, mean, and 90th percentile population exposures from
surface water systems are estimated to be near 0, 2.4, and 6.2 |ig/person/day, respectively. The
same values for populations exposed to bromoform from ground water systems are higher - 0.65,
3.8, and 9.6 |ig/person/day, respectively.
For purposes of comparison, estimates of ingestion exposure to bromodichloromethane,
dibromochloromethane, and bromoform in drinking water were also estimated from data collected
in other, older studies. Ingestion from ground water supplies was estimated from the median
levels found in the Ground Water Supply Survey conducted by U.S. EPA in 1980-81. Based on
the range of median levels (1.4-2.1 |ig/L (ppb)) and a consumption rate of two liters per day, the
median ingestion exposure to bromodichloromethane may range from 2.8 to 4.2 |ig/day.
Similarly, median exposure to dibromochloromethane may range from 4.2 to 7.8 |ig/day, and for
bromoform, median exposure may range from 4.8 to 8.4 |ig/day. Exposure to
bromodichloromethane from surface water supplies can be estimated based on the range of
median values observed under different conditions in the National Organics Monitoring Survey
conducted by U.S. EPA in 1976-1977, which mainly sampled surface water systems. Based on a
range of 5.9-14 |ig/L (ppb), exposure to bromodichloromethane from surface water is estimated
to be between 12 and 28 |ig/day. Similarly, based on the range of medians reported for
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dibromochloromethane concentrations, the median exposure is estimated to be up to 6 |ig/day.
The median levels of bromoform in the surface water supplies have been found to be less than the
EPA Drinking Water minimum reporting levels (MRLs) of 0.5-1 |ig/L (ppb). An estimate of
exposure based on the MRLs will be overly conservative because the actual concentration of
bromoform is not detectable. Based on the range of MRLs, 0.5-1 |ig/L (ppb), the exposure to
bromoform is estimated to range from 1 to 2 |ig/day for surface water supplies.
Ingestion exposure to brominated trihalomethanes in drinking water can also be estimated
from the concentrations found at the tap in the U.S. EPA's Total Exposure Assessment
Methodology (TEAM) study. Estimates of the average of the population intakes for ingestion of
bromodichloromethane from drinking water range from 0.42 to 42 /ig/person/day. The upper 90th
percentile estimates range from <2.0 to 90 //g/person/day. Estimates of the average population
intake of dibromochloromethane from drinking water range from 0.2 to 56 /ig/person/day. The
upper 90th percentile estimates range from < 0.9 to 86 //g/person/day. Estimates of the average of
the population intakes of bromoform, for those areas in which bromoform was measurable in a
majority of the samples, range from 1.6 to 16.2 //g/person/day. The upper 90th percentile
estimates range from 2.4 to 26 /ig/person/day. Four of the six locations in the TEAM study,
however, had a low frequency (less than 10%) of detection of bromoform in measurable
quantities.
Sources of uncertainty in these estimates of ingestion exposure include use of different
analytical methods, failure to report quantitation limits, using measures near the detection limit,
failure to report how nondetects are handled when averaging values (e.g., set to zero or one half
the detection limit), and failure to report sample storage method and duration. In addition, many
environmental factors influence the concentrations of these compounds in drinking water at the
tap and in vended or bottled waters used for drinking. These factors include season and
temperature, geographic location, source of water, residence time in distribution system, and
others.
The RSC (relative source contribution) is the percentage of total daily exposure that is
attributable to tap water when all potential sources are considered (e.g., air, food, soil, and
water). Ideally, the RSC is determined quantitatively using nationwide, central tendency and/or
high-end estimates of exposure from each relevant medium. In the absence of such data, a default
RSC ranging from 20% to 80% may be used.
The RSC used in the current and previous drinking water regulations for
dibromochloromethane is 80%. This value was established by use of a screening level approach
to estimate and compare exposure to dibromochloromethane from various sources. Information
considered for during this process is summarized in Appendix C. There are some uncertainties in
the 80%) RSC that are related to the availability of adequate concentration data for
dibromochloromethane in media other than water. Parallel RSC calculations were not performed
for bromodichloromethane and bromoform. The EPA has set the regulatory level for these
chemicals in drinking water at zero because it has been determined that they are probable human
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carcinogens. Therefore, determination of an RSC is not relevant for these chemicals because it is
the Agency's policy to perform RSC analysis only for noncarcinogens.
Brominated trihalomethanes have been detected in the blood and breast milk of humans.
A national survey of volatile organic compounds in whole blood detected bromodichloro-
methane, dibromochloromethane, and bromoform in 14%, 12%, and less than 10% of samples,
respectively, when highly sensitive analytical methods were applied. Several studies have
demonstrated that the level of individual brominated trihalomethanes in blood or breath increases
shortly after exposure to these compounds in tap water during bathing and showering. Exposure
during these events may occur by ingestion, dermal contact and/or inhalation of the volatilized
compound. In studies which examined households with differing concentrations of brominated
trihalomethanes in tap water, the levels of individual brominated trihalomethanes in blood or
exhaled breath paralleled the tap water concentration. The studies of showering and bathing
indicate that water use patterns and water quality characteristics are important variables in
determining the blood levels of brominated trihalomethanes. Dibromochloromethane was
detected in one of eight samples of breast milk collected from women living in the vicinity of U.S.
chemical manufacturing plants or user facilities.
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V. HEALTH EFFECTS IN ANIMALS
A. Acute Exposures
This section presents data on the acute effects of brominated trihalomethanes. Acute
lethality values for the brominated trihalomethanes are summarized in Table V-l. Additional
acute toxicity data are summarized in Table V-2.
1. Bromodichloromethane
Acute lethality of bromodichloromethane has been investigated in mice and rats. LD50
values for male and female ICR Swiss mice were 450 and 900 mg/kg, respectively (Bowman et
al., 1978). Chu et al. (1980) determined LD50 values of 916 and 969 mg/kg for male and female
Sprague-Dawley rats, respectively.
Bowman et al. (1978) administered bromodichloromethane in a single gavage dose in
Emulphor®:alcohol:saline (1:1:8) to ICR Swiss mice (10/sex/group). The administered doses
ranged from 500 to 4,000 mg/kg (individual doses not reported). Sedation and anesthesia
occurred at 500 mg/kg. Males were more sensitive to the lethal effects of bromodichloromethane
than females.
Table V-l Summary of LD50 Values for Brominated Trihalomethanes
Compound
LDS0 Values (mg/kg)
ICR Swiss Mouse a
Sprague-Dawley Rat b
Male
Female
Male
Female
Bromodichloromethane
450
900
916
969
Dibromochloromethane
800
1200
1186
848
Bromoform
1400
1550
1388
1147
Bowman et al. (1978)
bChu et al. (1980)
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February 20, 2003
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Table V-2 Summary of Acute Toxicity Studies for Brominated Trihalomethanes
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromodichloromethane
Bowman et al.
(1978)
Mouse
ICR
Swiss
Gavage
(aqueous)
M,
F
10
Single
dose
500 - 4000
Anesthesia, sedation at 500
mg/kg
NTP (1987)
Rat
F344/N
Gavage
(corn oil)
M,
F
5
Single
dose
150
300
600
1,250
2,500
Increased mortality at 600
mg/kg-day. Lethargy, labored
breathing at 1250 mg/kg-day and
above, 100% mortality at the
two highest dose groups
NTP (1987)
Mouse
B6C3F;
Gavage
(oil)
M,
F
5
Single
dose
150
300
600
1,250
2,500
100% mortality at the two
highest dose groups
Lilly et al.
(1994)
Rat
F344
Gavage
(corn oil)
(aqueous)
M
?
Single
dose
0
200 (LOAEL)
400
Renal tubule degeneration and
necrosis; alteration in markers of
renal function
Lilly et al.
(1996)
Rat
F344
Gavage
(corn oil
or water)
M
6
Single
dose
0
200 (LOAEL)
400
Renal tubule necrosis; alteration
in markers of renal function
Lilly et al.
(1997)
Rat
F344
Gavage
(aqueous)
M
5
Single
dose
0
123
164 (NOAEL)
246 (LOAEL)
328
492
Decreased body weight; elevated
liver and renal markers
Keegan et al.
(1998)
Rat
F344
Gavage
(aqueous)
M
6
Single
dose
0
21
31
41 (NOAEL)
82 (LOAEL)
123
164
246
Elevated renal markers;
decreased liver weight and body
weight
Dibromochloromethane
Bowman et al.
(1978)
Mouse
ICR
Swiss
Gavage
(aqueous)
M,
F
10
Single
dose
500 - 4000
Anesthesia, sedation
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Table V-2 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
NTP (1985)
Rat
F344/N
Gavage
(corn oil)
M,
F
5
Single
dose
160
310
630
1250
2,500
Increased mortality at 630 mg/kg
and above, with 100% mortality
in high dose group
NTP (1985)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
5
Single
dose
160
310
630
1250
2500
Increased mortality in females at
630 mg/kg-day and above and in
males at 310 mg/kg-day and
above; 100% mortality in males
in two highest dose groups and
females in two highest dose
groups
Mtiller et al.
(1997)
Rat
Wistar
Gavage
(olive oil)
M
6
Single
dose
0
83
167
333
667
Transient decrease in blood
pressure and heart rate;
decreased activity; effects on
heart muscle contractility and
changes in some cardiac
parameters
Bromoform
Bowman et al.
(1978)
Mouse
ICR
Swiss
Gavage
(aqueous)
M,
F
10
Single
dose
500 - 4000
Ataxia, sedation, and anesthesia
at 500 mg/kg
Chu et al.
(1980)
Rat
SD*
Gavage
(corn oil)
M,
F
10
Single
dose
546
765
1071
1500
2100
Sedation, ataxia, liver and
kidney congestion
NTP (1989a)
Rat
F344/N
Gavage
(corn oil)
M,
F
5
Single
dose
125
250
500
1,000
2000
No deaths at 500 and lower;
60% mortality at 1,000; 100%
mortality at 2,000;
shallow breathing in two highest
dose groups
NTP (1989a)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
5
Single
dose
125
250
500
1,000
2000
10%) mortality at 500 mg/kg-day
* SD, Sprague-Dawley
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NTP (1987) administered single gavage doses of bromodichloromethane in corn oil to
male and female F344/N rats and B6C3FX mice (5/sex/dose) at 150, 300, 600, 1,250, or
2,500 mg/kg. Animals were observed for 14 days, and a necropsy was performed on at least one
male and one female in each dose group. All animals dosed with 1,250 or 2,500 mg/kg died
before the end of the study. At 600 mg/kg, deaths occurred in two of five male rats, one of five
female rats, five of five male mice, and two of five female mice. Clinical signs observed in rats at
1,250 or 2,500 mg/kg included lethargy and labored breathing. Clinical signs observed in mice at
or above 600 mg/kg included lethargy, with the exception that this sign was not observed in high-
dose male mice. At necropsy, the liver from animals dosed with 1,250 or 2,500 mg/kg appeared
pale. No dose-related effects were seen on body weight gain in animals that survived.
Lilly et al. (1994) examined the effect of vehicle on the toxicity of bromodichloro-
methane. Male F344 rats (number unknown) were administered a single dose of 0, 200, or 400
mg/kg bromodichloromethane by gavage in corn oil or in an aqueous 10% Emulphor® solution.
Body weights were significantly decreased at 400 mg/kg only in the animals receiving the aqueous
gavage. Absolute and relative kidney weights were significantly increased at 400 mg/kg in both
vehicles, with a significantly greater increase observed in the animals gavaged with oil compared
to those gavaged with 10% Emulphor® solution. Serum markers of hepatotoxicity were
significantly increased at 400 mg/kg in both vehicles with one nonsignificant increase in the
aqueous vehicle. The increases were significantly greater for two of these markers in animals
receiving the oil vehicle compared to those receiving the aqueous vehicle. Clinical observations
were supported by histopathology findings. Hepatocellular degeneration and necrosis were
observed at 400 mg/kg in both vehicles. The difference in vehicles was reflected in more severe
hepatocellular degeneration and a higher incidence of centrilobular necrosis in animals receiving
the oil gavage compared to those receiving the aqueous gavage. Numerous increases in urinary
markers of renal toxicity were observed 24 hours after dosing. Based on the differences
observed, renal toxicity at 200 mg/kg was similar or greater in the aqueous vehicle. Renal toxicity
at 400 mg/kg, however, was greater in the oil vehicle. The time to peak toxicity was both dose-
and vehicle-dependent. At 200 mg/kg, peak damage was observed at 24 hours in both vehicles.
At 400 mg/kg, peak damage was observed at 48 hours following oil gavage and at 24-36 hours
following aqueous gavage. Histopathology revealed both renal tubule degeneration and necrosis
at both dose levels. The incidence of renal tubule degeneration was greater in animals receiving
the aqueous gavage at the low dose; however, the severity of renal degeneration and necrosis was
greater in the animals receiving the oil gavage at the high dose. The authors attributed the vehicle
differences to slower gastrointestinal uptake of bromodichloromethane from the oil vehicle
compared to the aqueous vehicle. At the high dose, more bromodichloromethane would be
converted to a reactive metabolite following oil dosing, while saturation would occur following
aqueous dosing. At the low dose, the difference in uptake would have less of an effect. Overall,
this study found that the kidney was more sensitive than the liver to a single dose of
bromodichloromethane. A LOAEL of 200 mg/kg was identified for each vehicle based on
minimal renal tubule degeneration and changes in markers of renal function.
Lilly et al. (1996) investigated the effect of subchronic pretreatment with corn oil on the
toxicity of bromodichloromethane. Prior to initiation of dosing with bromodichloromethane, male
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Fischer 344 rats (6 animals/group) were gavaged with oral doses of corn oil or water for six
weeks (5 days/week) at a constant volume of 5 mL/kg. Following pretreatment, the animals
were gavaged with a single dose of 0, 200, or 400 mg bromodichloromethane/kg in 10%
Emulphor®. Urine was collected at 24, 36, and 48 hours following bromodichloromethane
administration. The rats were sacrificed at 48 hours and necropsies were performed. Activities of
the hepatotoxicity indicators alanine aminotransferase (ALT), aspartate aminotransferase (AST),
lactate dehydrogenase (LDH), and sorbitol dehydrogenase (SDH) were measured in the serum,
and the renal toxicity indicators alkaline phosphatase (ALK), AST, and LDH, were measured in
the urine. Additional analyses included determination of serum levels of bile acids, triglycerides,
cholesterol and albumin, and urine levels of N-acetylglucosaminidase and gamma glutamyl
transpeptidase activity. Enzymatic activity of cytochrome P450 isoforms CYP2E1 and
CYP2B1/B2 was measured in the microsomal fraction of the liver to investigate whether corn oil
was an inducer of bromodichloromethane metabolizing enzymes.
Liver weight was significantly reduced only in the water pretreatment group at the high
dose. Kidney weight was reduced in both pretreatment groups at the high dose. Activities of
serum AST and LDH were significantly elevated in both pretreatment groups at 400 mg/kg. ALT
levels increased in a dose-dependent manner in the water pretreatment group, but significant
elevations were noted only at the 400 mg/kg dose in animals pretreated with corn oil. Activities
of urinary AST and LDH were greater than controls in both pretreatment groups after 24, 36, and
48 hours. ALK levels were significantly increased in both pretreatment groups at 24 hours. At
36 and 48 hours, ALK levels were elevated only in water-pretreated animals. High incidences of
renal tubular necrosis occurred at 200 and 400 mg/kg in both pretreatment groups. There were
no significant differences in the histopathological lesion scores between the pretreatment groups.
No significant differences were noted in the hepatic activity of CYP2E1 or CYP2B1/B2 in the
corn oil pretreated animals compared to the water controls. Although a number of differences
between the pretreatment groups were noted in results for specific endpoints, the overall results
from this study indicate that 6 weeks of pretreatment with corn oil did not significantly enhance
the acute hepato- or nephrotoxicity of bromodichloromethane. In addition, the reported data
suggest that vehicle-related differences in toxicity observed in other bromodichloromethane
studies are most likely due to pharmacokinetic differences in absorption rather than altered
enzyme activity induced by corn oil. This study confirms the acute LOAEL of 200 mg/kg-day
previously identified by Lilly et al. (1994) for renal toxicity.
Lilly et al. (1997) administered single doses of bromodichloromethane by gavage in
aqueous 10% Emulphor® solution to male F344 rats at dose levels of 0, 123, 164, 246, 328, or
492 mg/kg. Groups of 5 animals/dose were sacrificed at 24 and 48 hours post-dosing. Body
weights were significantly decreased at or above 246 mg/kg after 48 hours. At 24 hours, absolute
and relative kidney weights were significantly increased at or above 328 mg/kg and 246 mg/kg,
respectively. At 48 hours, only relative kidney weight at the high dose was significantly increased.
At 24 hours, serum markers of liver damage (ALT and AST) were significantly increased at or
above 246 mg/kg with one marker (SDH) increased at all dose levels. Although smaller
statistically significant increases were observed at the low doses at 24 hours for ALT (123 and
164 mg/kg) and AST (164 mg/kg), the biological significance of these increases is unclear. After
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48 hours, serum levels of these markers were decreased from 24-hour levels with statistically
significant changes noted only at the higher doses. No effects in urinary markers of kidney
damage were found at either 123 or 164 mg/kg. These markers, however, were significantly
elevated after 24 hours for doses at or above 246 mg/kg with few exceptions. No
histopathological examination was conducted. These results were generally consistent with
earlier results (Lilly et al., 1994), although the present study was conducted at doses low enough
to identify a NOAEL. In contrast to the earlier results of Lilly et al. (1994), this study did not
find that the kidney was more sensitive than the liver to the toxic effects of
bromodichloromethane. Based on hepatotoxicity and nephrotoxicity, this study identified a
NOAEL of 164 mg/kg and a LOAEL of 246 mg/kg.
Keegan et al. (1998) investigated the acute toxicity of bromodichloromethane
administered orally in an aqueous vehicle. Male Fischer 344 rats (6 animals/group) were gavaged
with a single dose of 0, 0.125, 0.1875, 0.250, 0.5, 0.75, 1.0 or 1.5 mmol/kg dissolved in a 10%
aqueous solution of Alkamuls EL-620. These doses of bromodichloromethane are equivalent to
0, 20.5, 30.7, 41.0, 81.9, 122.9, 163.8, and 245.7 mg/kg, respectively. Control animals were
dosed with vehicle only (10% Alkamuls EL-620). Gavage volumes were kept constant at 5 ml/kg
body weight. Animals were sacrificed 24 hours after dose administration and the liver, kidneys,
and serum were harvested. Significant decreases in body weight were observed in the 0.75, 1.0,
and 1.5 mmol bromodichloromethane/kg treated animals. Decreases in absolute liver weights
were observed in the 0.5, 0.75, 1.0, and 1.5 mmol bromodichloromethane/kg treated animals. No
change was noted in relative liver weights. Absolute kidney weights were not effected by
bromodichloromethane treatment, but relative kidney weights were significantly increased in the
two highest dose groups (1.0 and 1.5 mmol bromodichloromethane/kg). Serum levels of ALT,
SDH, and AST were assessed as an indication of liver toxicity. Dose-dependent elevations in
ALT (45%) to 239% increase), AST (25%> to 130%) increase) and SDH (74%> to 378%> increase)
were observed in the 0.5, 0.75, 1.0, and 1.5 mmol dose groups. Based on these findings, 0.25
mmol/kg (41.0 mg/kg) represents the NOAEL and 0.5 mmol/kg (81.9 mg/kg) represents the
LOAEL for orally administered bromodichloromethane in an aqueous vehicle. These authors
used the NOAEL of 41.0 mg/kg to calculate One-Day Health Advisories for drinking water of 4
mg/L for a 10-kg child and 14 mg/L for a 70-kg adult.
2. Dibromochloromethane
The acute oral lethality of dibromochloromethane has been assessed in rats and mice of
both sexes. Chu et al. (1980) reported LD50 values in male and female Sprague-Dawley rats of
1,186 and 848 mg/kg-day for males and females, respectively. Bowman et al. (1978) reported
LD50 values in mice of 800 and 1,200 mg/kg-day, respectively.
Bowman et al. (1978) investigated the acute oral toxicity of dibromochloromethane in
ICR Swiss mice (10/sex/group). Doses of 500 to 4000 mg/kg (individual doses not reported)
were administered by gavage in Emulphor®:alcohol:saline (1:1:8) to fasted animals. Sedation and
anesthesia occurred at 500 mg/kg. Males were more sensitive than females to the acute lethal
effects of dibromochloromethane.
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NTP (1985) evaluated the acute toxicity of dibromochloromethane in male and female
F344/N rats. The rats (5 animals/sex/dose) received single doses of 160, 310, 630, 1,250, or
2,500 mg/kg dibromochloromethane by gavage in corn oil. The observation period following
treatment was 14 days. Mortality in high-dose rats was 100% by day 3. At the 1,250 mg/kg
dose, four male rats and one female rat died. One female rat died in the 630 mg/kg group. Doses
of 310 mg/kg or greater produced lethargy in all animals for 3 hours after dosing. A gross
necropsy was conducted on one or two animals from each group. No treatment-related effects
were observed in rats selected for gross necropsy.
In a concurrent study, NTP (1985) evaluated the acute toxicity of dibromochloromethane
in male and female B6C3FX mice (5/sex/dose). The mice received single doses of 160, 310, 630,
1,250, or 2,500 mg/kg dibromochloromethane by gavage in corn oil. The observation period
following treatment was 14 days. All male mice receiving the 2,500 mg/kg and 1,250 mg/kg
doses died. Three male mice receiving the 630 mg/kg dose died, while a single male mouse died
at the 310 mg/kg dose. All female mice receiving the 2500 mg/kg dose died. Four of the female
mice administered the 1,250 mg/kg dose died between days 2 and 8 post-treatment. No female
mice died at doses of 630 mg/kg or lower. A gross necropsy was conducted on one or two
animals from each group. At necropsy, aberrations of the kidney (dark red or pale medullae) and
liver (discolored foci) were reported to be more frequently observed in treated animals than in
control animals (raw data were not presented in the study).
Miiller et al. (1997) investigated the cardiotoxic effects of acute dibromochloromethane
exposure. Male Wistar rats were administered a single dose of dibromochloromethane by gavage
in olive oil at dose levels of 0, 83, 167, 333, or 667 mg/kg. Telemetric measurements of
cardiovascular parameters (heart rate, blood pressure, body temperature, and physical activity)
were recorded in conscious rats (6/group) 24 hours prior to administration to 72 hours following
administration. Heart rate and blood pressure were also measured in urethane-anesthetized rats
(10/group) 25 minutes following administration. For these rats, contractility parameters, such as
the Krayenbiihl index, were also calculated. Treatment-related arrhythmias were not observed in
conscious rats dosed with 83 to 333 mg/kg of dibromochloromethane while rats in the high-dose
group exhibited premature ventricular contractions one minute following administration. Heart
rate and body temperature were initially decreased in all treatment groups following
administration, but returned to control values 24 hours post-exposure in rats administered 83 to
333 mg/kg. In the high-dose rats, heart rate remained depressed up to 48 hours post-exposure,
and body temperature decreased 4.5°C below control values by 72 hours post-exposure. Blood
pressure was initially increased in all treatment groups following administration, but began to
return to control values within 48 hours post-exposure in rats administered 83 to 333 mg/kg.
Blood pressure in the high-dose group, however, decreased below control values 72 hours post-
exposure. Physical activity was decreased in conscious rats administered 333 and 667 mg/kg
during the entire observation period. In urethane-anesthetized rats, negative effects on muscle
contractility were observed at dose levels of 333 and 667 mg/kg, negative chronotropic (rate of
contraction) effects were observed at the 333 mg/kg dose level, and negative dromotropic
(defined as influencing the velocity of conduction of excitation, as in nerve or cardiac muscle
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fibers) effects were observed at dose levels of 167 to 667 mg/kg. Heart rate, blood pressure, and
several contractility parameters, however, did not exhibit dose-related trends.
3. Bromoform
Bowman et al. (1978) assessed the acute oral toxicity of bromoform in ICR Swiss mice.
Groups of ten male (30 to 35 g) and ten female (25 to 30 g) mice were treated with single doses
ranging from 500 to 4,000 mg/kg. Compounds were solubilized in Emulphor®:alcohol:saline
(1:1:8) and administered by gavage to fasted animals. The period of observation following
treatment was 14 days. LD50 values were 1400 and 1550 mg/kg for males and females,
respectively. Ataxia, sedation, and anesthesia occurred within 60 minutes of treatment at doses of
1000-mg/kg and above. Sedation lasted approximately 4 hours.
Chu et al. (1980) evaluated the acute toxicity of bromoform in male and female Sprague-
Dawley rats. Fasted adult rats (10/sex/dose) received doses of 546, 765, 1071, 1500, or 2100
mg/kg bromoform dissolved in corn oil by gavage. Clinical observations were made for 14 days
after treatment. The LD50 values for male and female rats were 1388 and 1147 mg/kg,
respectively. Clinical signs observed in treated rats included sedation, flaccid muscle tone, ataxia,
piloerection, and hypothermia. Gross pathological examination revealed liver and kidney
congestion in treated animals. Chu et al. (1982a) reported results for growth, food intake, organ
weight, histopathology, hematological indices, liver microsome aniline hydroxylase activity and
serum chemistry in surviving rats. Bromoform treatment increased liver protein concentration in
the serum of male rats at doses of 765 and 1071 mg/kg. Lymphocyte counts were decreased in
male (765 and 1071 mg/kg doses) and female (765 mg/kg) rats but the effect was not dose-
dependent. Female rats at the 765 mg/kg dose had elevated aniline hydroxylase levels.
NTP (1989a) investigated the acute oral toxicity of bromoform in male and female
F344/N rats. The rats (5/sex/group) were administered a single oral dose of bromoform (by
gavage, in corn oil) at dose levels of 125, 250, 500, 1,000, or 2,000 mg/kg. Control groups were
not included in the study design. Mortality was 10/10 at 2,000 mg/kg, 6/10 at 1,000 mg/kg, and
0/10 at 500 mg/kg or lower. Shallow breathing was observed in rats that received the 1000 or
2000 mg/kg doses. No other clinical signs were reported.
NTP (1989a) investigated the acute oral toxicity of bromoform in male and female
B6C3F, mice. The mice received a single oral dose of bromoform (by gavage, in corn oil) at dose
levels of 125, 250, 500, 1,000, or 2,000 mg/kg. There were no controls. Mortality was 0/10 at
2,000 mg/kg, 6/10 at 1,000 mg/kg, 1/10 at 500 mg/kg, and 0/10 at 250 mg/kg or lower. The final
mean body weight of mice that survived to the end of the study period was unaffected by
bromoform exposure. Male mice that received doses of 500, 1,000, or 2,000 mg/kg and females
that received 1,000 or 2,000 mg/kg were lethargic. Shallow breathing was noted in male mice
administered the 1,000 or 2,000 mg/kg dose.
B. Short-Term Exposures
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This section summarizes short-term studies (less than approximately 90 days) on the
health effects of brominated trihalomethanes in animals. Details of these studies are summarized
in Table V-3.
Table V-3 Summary of Short Term Toxicity Studies for Brominated Trihalomethanes
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromodichloromethane
Oral Exposure
Chu et al.
(1982a)
Rat
SD*
Drinking
water
M
10
28 days
0
0.8
8
68 (NOAEL)
No signs of toxicity observed.
Munson et al.
(1982)
Mouse
CD-I
Gavage
(aqueous)
M,
F
8-12
14 days
0
50 (NOAEL)
125 (LOAEL)
250
Decreased immune function;
increased liver weight,
decreased absolute and relative
spleen wt. (females)
Condie et al.
(1983)
Mouse
CD-I
Gavage
(corn oil)
M
8-16
14 days
0
37
74 (NOAEL)
148 (LOAEL)
Liver and kidney
histopathology
NTP (1987)
Rat
F344/N
Gavage
(corn oil)
M,
F
5
14 days
0
38
75
150 (NOAEL)
300 (LOAEL)
600
Decreased body weight gain;
renal pathology
NTP (1987)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
5
14 days
0
19
38
75 (NOAEL)
150 (LOAEL)
300
Mortality, renal histopathology
Aida et al.
(1992a)
Rat
Wistar
Diet
M
7
1 month
0
21
62 (NOAEL)
189 (LOAEL)
Liver histopathology
Aida et al.
(1992a)
Rat
Wistar
Diet
F
7
1 month
0
21
66 (NOAEL)
204 (LOAEL)
Liver histopathology
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Table V-3 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Thornton-
Manning et al.
(1994)
Rat
F344
Gavage
(aqueous)
F
6
5 days
0
75 (NOAEL)
150 (LOAEL)
300
Liver histopathology, renal
histopathology; increased liver
and kidney wt.;
elevated markers of
hepatotoxicity
Thornton-
Manning et al.
(1994)
Mouse
C57BL/
63
Gavage
(aqueous)
F
6
5 days
0
75 (NOAEL)
150 (LOAEL)
Increased serum markers of
hepatotoxicity
Potter et al.
(1996)
Rat
F344
Gavage
(aqueous)
M
1,3, or 7
days
123
246 (NOAEL)
No effect in hyaline droplet
formation or cell proliferation
Melnick et al.
(1998)
Mouse
B6C3F;
Gavage
(corn oil)
F
10
3 weeks
(5 d/wk)
0
75 (NOAEL)
150 (LOAEL)
326
Increased abs. and relative liver
weight; increased serum
markers of hepatotoxicity;
hepatocyte degeneration;
increased labeling index
NTP (1998)
Rat
SD
Drinking
water
M,
F
6
2 weeks
0
11
45 (NOAEL)
91 (LOAEL)
124
Transient reduction in weight
gain
NTP (1998)
Rat
SD
Drinking
water
M,
F
5-13
35 days
Group A males
0
9 (NOAEL)
38 (LOAEL)
67
Single cell hepatic necrosis in
Group A males
Coffin et al.
(2000)
Mouse
B6C3F;
Gavage
(Corn oil)
Drinking
water
F
10
11 days
11 days
0
150 (LOAEL)
300
0
138 (LOAEL)
Hydropic degeneration in liver
(corn oil gavage and drinking
water); increased relative liver
weight (gavage); increased
proliferating cell nuclear
antigen labeling index (gavage)
Inhalation Exposure
Torti et al.
(2001)
Mouse
C57BL/6
FVB/N
(wild-
type)
Vapor
M
6
1 week
(6 hr/day)
0 ppm
1 ppm
(NOAEL)
10 ppm
(LOAEL)
30 ppm
100 ppm
150 ppm
Dose-dependent marginal
increase in renal tubular
degeneration in C57BL/6 mice;
mild increase in renal tubular
degeneration and marginal
increase in hepatic
degeneration in FVB/N mice;
sign, increased labeling index
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Table V-3 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Torti et al.
(2001)
Mouse
C57BL/6
(p53
hetero-
zygous)
Vapor
M
6
1 week
(6 hr/day)
0
1 ppm
(NOAEL)
10 ppm
(LOAEL)
30 ppm
100 ppm
150 ppm
Dose-dependent marginal to
mild increase in renal tubular
degeneration; sign, increased
labeling index
Torti et al.
(2001)
FVB/N
(p53
hetero-
zygous)
Vapor
M
6
1 week
(6 hr/day)
0 ppm
0.3 ppm
1 ppm
3 ppm
(NOAEL)
10 ppm
(LOAEL)
30 ppm
Dose-dependent mild increase
in renal tubular degeneration
and marginal increase in
nephrosis; marginal increase in
hepatic degeneration; sign,
increased relative kidney wt.
and labeling index.
Torti et al.
(2001)
Mouse
C57BL/6
FVB/N
(wild-
type)
Vapor
M
6
3 weeks
(6 hr/day)
0 ppm
0.3 ppm
1 ppm
3 ppm
(NOAEL)
10 ppm
(LOAEL)
30 ppm
Marginal increase in renal
tubular degeneration
Torti et al.
(2001)
Mouse
C57BL/6
FVB/N
(p53
hetero-
zygous)
Vapor
M
6
3 weeks
(6 hr/day)
0 ppm
0.3 ppm
1 ppm
3 ppm
(NOAEL)
10 ppm
(LOAEL)
30 ppm
Marginal or mild increase in
renal tubular degeneration in
both strains; marginal increase
in hepatic degeneration in
FVB/N heterozygous strain
Dibromochloromethane
Munson et al.
(1982)
Mouse
CD-I
Gavage
(aqueous)
M,
F
8-12
14 days
0
50 (NOAEL)
125 (LOAEL)
250
Decreased immune function
Chu et al.
(1982a)
Rat
SD
Drinking
water
M
10
28 days
0
0.7
8.5
68 (NOAEL)
No effect on growth, clinical
signs, biochemical or
histopathological endpoints
Condie et al.
(1983)
Mouse
CD-I
Gavage
(corn oil)
M
8-16
14 days
0
37
74 (NOAEL)
147 (LOAEL)
Decreased PAH uptake,
moderate liver and kidney
histopathology
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Table V-3 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
NTP (1985)
Rat
F344/N
Gavage
(corn oil)
M,
F
5
14 days
0
60
125
250 (NOAEL)
500 (LOAEL)
1,000
Mortality; liver and renal gross
pathology
NTP (1985)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
5
14 days
0
30
60 (NOAEL)
125 (LOAEL)
250
500
Liver and kidney gross
pathology
Aida et al.
(1992a)
Rat
Wistar
Diet
M
7
1 month
0
18 (NOAEL)
56 (LOAEL)
173
Liver histopathology
Aida et al.
(1992a)
Rat
Wistar
Diet
F
7
1 month
0
34 (NOAEL)
101 (LOAEL)
333
Liver histopathology; increased
relative liver weight
Potter et al.
(1996)
Rat
F344
Gavage
(aqueous)
M
1,3, or 7
days
156
312 (NOAEL)
No effect on hyaline droplet
formation or cell proliferation
Melnick et al.
(1998)
Mouse
B6C3F;
Gavage
(corn oil)
F
10
3 weeks
(5 d/wk)
0
50
100 (NOAEL)
192 (LOAEL)
417
Liver histopathology; increased
serum enzymes and liver
weight
Coffin et al.
(2000)
Mouse
B6C3F;
Gavage
(Corn oil)
Drinking
water
F
10
11 days
11 days
0
100 (LOAEL)
300
0
171
Increased proliferating cell
nuclear antigen labeling index
(gavage); increased relative
liver wt.
Bromoform
Munson et al.
(1982)
Mouse
CD-I
Gavage
(aqueous)
M,
F
6-12
14 days
0
50
125 (NOAEL)
250 (LOAEL)
Increased serum enzyme
activity (AST); decrease in
antibody forming cells and
delayed-type hypersensitivity
response
Chu et al.
(1982a)
Rat
SD
Drinking
water
M
10
28 days
0.7
8.5
80 (NOAEL)
No effect on growth, clinical
signs, biochemical or
histopathological endpoints
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Table V-3 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Condie et al.
(1983)
Mouse
CD-I
Gavage
(corn oil)
M
8-16
14 days
0
72
145 (NOAEL)
289 (LOAEL)
Decreased PAH uptake,
moderate histopathological
changes
NTP (1989a)
Rat
F344/N
Gavage
(corn oil)
M,
F
5
14 days
0
100
200 (NOAEL)
400 (LOAEL)
600
800
Decreased body weight gain;
1/5 died at 400 mg/kg-
day; 100% mortality at two
highest doses
NTP (1989a)
Mouse
B6C3F;
Gavage
(corn oil)
M
5
14 days
0
50
100
200 (NOAEL)
400 (LOAEL)
600
Stomach nodules;
ataxia, lethargy; 1/5 died at
high dose
Aida et al.
(1992a)
Rat
Wistar
Diet
M
7
1 month
0
62 (NOAEL)
187 (LOAEL)
618
Hepatic vacuolization, serum
chemistry/biochemistry
Aida et al.
(1992a)
Rat
Wistar
Diet
F
7
1 month
0
56 (NOAEL)
208 (LOAEL)
728
Hepatic vacuolization, serum
chemistry/biochemistry
Potter et al.
(1996)
Rat
F344
Gavage
(aqueous)
M
1,3, or 7
days
190
379 (NOAEL)
No effect on hyaline droplet
formation or cell proliferation
Melnick et al.
(1998)
Mouse
B6C3F;
Gavage
(corn oil)
F
10
3 weeks
(5 d/wk)
0
200 (NOAEL)
500 (LOAEL)
Increase in absolute and
relative liver wt.; marginally
significant increase in LI at
highest dose
Coffin et al.
(2000)
Mouse
B6C3F;
Gavage
(corn oil)
Drinking
Water
F
10
11 days
11 days
0
200 (LOAEL)
500
0
301
Liver histopathology; increased
proliferating cell nuclear
antigen labeling index;
increased relative liver wt.
* SD, Sprague-Dawley
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1. Bromodichloromethane
Munson et al. (1982) administered bromodichloromethane by aqueous gavage to male and
female CD-I mice (8 to 12/sex/group) for 14 days at levels of 0, 50, 125, or 250 mg/kg-day.
Endpoints evaluated included body and organ weights, hematology, serum enzyme levels (SGOT,
SGPT), and humoral and cell-mediated immune system functions. At 250 mg/kg-day, body
weights were significantly decreased. Significant organ weight changes included increased
relative liver weight (mid- and high-dose groups), decreased absolute spleen weight (high-dose
males and mid- and high-dose females), and decreased relative spleen weight (mid- and high-dose
females).
Among the hematology endpoints, only fibrinogen levels were significantly decreased in
high- dose males and in mid- and high-dose females. Significant clinical chemistry findings
included decreased glucose levels (high-dose males), increased ALT and AST activities (high-
dose groups), and increased blood urea nitrogen (BUN) levels (high-dose groups).
Bromodichloromethane appeared to affect the humoral immune system, as judged by significantly
decreased antibody-forming cells (high-dose males and mid- and high-dose females) and
hemagglutination titers (mid- and high-dose males and high-dose females). This study identified a
NOAEL of 50 mg/kg-day and a LOAEL of 125 mg/kg-day for bromodichloromethane on the
basis of decreased immune function in females.
Chu et al. (1982a) administered bromodichloromethane to male Sprague-Dawley rats
(10/group) in drinking water for 28 days at dose levels of 0, 5, 50, or 500 ppm. These levels
corresponded to doses of 0, 0.8, 8.0, or 68 mg/kg-day, as calculated by the authors based on
recorded fluid intake. The authors observed no effects on growth rate or food consumption and
no signs of toxicity throughout the exposure. No dose-related biochemical or histologic changes
were detected (no data were provided). This study identified a NOAEL of 68 mg/kg-day, but the
reported data were too limited to allow an independent verification.
Condie et al. (1983) investigated the renal and hepatic toxicity of bromodichloromethane
in male CD-I mice (8 to 16/group). Bromodichloromethane was administered by gavage in corn
oil for 14 days at dose levels of 0, 37, 74 or 148 mg/kg-day. Biochemical evidence of liver
damage (significantly elevated ALT) was observed at the high dose, while biochemical evidence
of kidney damage (significantly decreased p-aminohippurate (PAH) uptake by kidney slices) was
observed at the mid and high dose. Significantly decreased BUN levels were observed in the low-
and mid-dose groups, but not in the high-dose group. Histopathology revealed no consistent or
important changes at the low or mid-level doses, with minimal to moderate liver and kidney injury
observed in the majority of animals at the high dose. Liver lesions included centrilobular pallor
and focal inflammation. Kidney lesions included intratubular mineralization, epithelial hyperplasia,
and cytomegaly. Although the severity of these lesions was primarily minimal to slight, a few
animals in the high dose group exhibited moderate to moderately severe intratubular
mineralization and/or epithelial hyperplasia. This study identified a NOAEL value of 74 mg/kg-
day and a LOAEL value of 148 mg/kg-day for bromodichloromethane, based on histopathology.
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NTP (1987) administered doses of 0, 38, 75, 150, 300, or 600 mg/kg-day of
bromodichloromethane in corn oil by gavage to male and female F344/N rats (5/sex/dose) for
14 days. One low-dose and one high-dose female died before study termination. All high-dose
animals were hyperactive after dosing and either lost weight or gained no weight during the study.
Final mean body weights were not significantly affected in groups given 38, 75, or 150 mg/kg-
day. At 300 mg/kg, body weights of males and females were decreased by 21% and 7%,
respectively, relative to vehicle controls. At 600 mg/kg-day, body weights of males and females
were decreased by 44% and 22%, respectively, relative to vehicle controls. Necropsy was
performed on all animals. Renal medullae were reddened in all high-dose males and in one female
in each of the control, low-dose, and high-dose groups. This study identified a NOAEL of
150 mg/kg-day and a LOAEL of 300 mg/kg-day in rats, based on decreased body weight gain.
In a parallel experiment, NTP (1987) administered doses of 0, 19, 38, 75, 150, or
300 mg/kg-day bromodichloromethane in corn oil by gavage to male and female B6C3F, mice
(5/sex/dose) for 14 days. All male mice that received 150 or 300 mg/kg-day
bromodichloromethane died before study termination. Clinical signs included lethargy,
dehydration, and hunched posture. The final mean body weights of the mice that survived were
not significantly different from the controls. The renal medullae were reddened in four males in
the 150 mg/kg-day group, all males in the 300 mg/kg-day group, and one female in the 150
mg/kg-day group. Based on behavior, appearance, gross necropsy, and mortality, this study
identified a NOAEL of 75 mg/kg-day and a frank effect level (FEL) of 150 mg/kg-day in male
mice. An interesting point to note is that this study and the study by Condie et al. (1983) were
conducted under similar conditions (mice administered bromodichloromethane by gavage in corn
oil for 14 days), but with dramatically different results. In contrast to the 100% mortality
observed in this study for male mice, Condie et al. (1983) found only moderate histopathology in
male CD-I mice at 148 mg/kg-day with no deaths occurring. The reason for this difference is
unclear, but may be related to strain-specific differences in sensitivity.
Aida et al. (1992a) administered bromodichloromethane to Slc:Wistar rats (7/sex/group)
for one month at dietary levels of 0%, 0.024%, 0.072%, or 0.215%) for males and 0%, 0.024%),
0.076%o, or 0.227%o for females. The test material was microencapsulated and mixed with
powdered feed; placebo granules were used for the control groups. Based on the mean food
intakes, the study authors reported the mean compound intakes for the one-month period as 0,
20.6, 61.7, or 189.0 mg/kg-day for males and 0, 21.1, 65.8, or 203.8 mg/kg-day for females.
Clinical effects, body weight, food consumption, hematology parameters, serum chemistry, and
histopathology of all major organs were determined. Body weights were significantly decreased
in the high-dose groups relative to the controls. The high-dose animals also exhibited slight
piloerection and emaciation. Relative liver weight was increased only in high-dose females.
Significant, dose-related biochemical findings at the low dose were limited to decreased LDH
levels in males, but the biological significance of this effect is unclear. Serum LDH levels were
also significantly decreased at the low and high dose in females. Other statistically significant,
dose-related changes included decreased glucose (high-dose males), decreased serum triglycerides
(high-dose groups), decreased serum cholinesterase activity (high-dose males and mid- and high-
dose females), and increased total cholesterol (mid- and high-dose males). The changes in
cholinesterase activity and cholesterol levels in males were not dose-related. The cholesterol
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levels were within normal ranges at all doses. Treatment-related histopathological lesions were
limited to the liver and were rated as very slight or slight. The lesions were mostly confined to the
high-dose groups. Vacuolization observed in mid-dose females and in a single low-dose male was
not considered an adverse effect. Other observed effects included swelling of hepatocytes, single
cell necrosis, hepatic cord irregularity, and bile duct proliferation. These lesions were observed
only in high-dose males and females with the exception of very slight to slight changes in
individual low-dose males. No effect was observed on any hematology parameter. Based on the
histopathology observed in high-dose males and females, the LOAELs identified in this study for
bromodichloromethane in rats were 189.0 mg/kg-day in males and 203.8 mg/kg-day in females;
the NOAELs were 61.7 mg/kg-day in males and 65.8 mg/kg-day in females.
Thornton-Manning et al. (1994) administered bromodichloromethane at dose levels of 0,
75, 150, or 300 mg/kg-day by gavage to female F344 rats (6 animals/dose) for five consecutive
days. The dosing vehicle consisted of an aqueous 10% Emulphor® solution. Animals were
sacrificed on day 6. Two animals in the high-dose group died on day 5. Final body weights of the
high-dose group were significantly decreased compared to the controls. Absolute and relative
kidney and liver weights were significantly increased at 150 and 300 mg/kg-day with the
exception of a nonsignificant increase in absolute liver weight at 150 mg/kg-day. Toxic effects on
the kidney and liver were reflected in significantly increased LDH, AST, SDH, creatinine, and
BUN at 300 mg/kg-day. These results were supported by the histopathology findings. In the
liver, centrilobular vacuolar degeneration was observed at both 150 and 300 mg/kg-day with the
severity of the effect increased with increasing dose. Centrilobular hepatocellular necrosis was
also observed in one high-dose animal. In the kidney, renal tubular vacuolar degeneration and
renal tubule regeneration were observed at 150 and 300 mg/kg-day with the incidence and
severity increased with increasing dose. While minimal renal tubule necrosis was observed in only
one animal at the mid dose, all animals at the high dose exhibited mild to moderate renal tubule
necrosis. Significant decreases in the hepatic activity of the CYP1A and CYP2B markers
ethoxyresorufin-O-dealkylase (EROD) and pentoxyresorufin-O-dealkylase (PROD), were
observed at all doses. The effect, however, was not dose-related. No effect on the CYP2E1
marker pNP-hydroxylase, was observed. Based on kidney and liver lesions observed at the mid
dose, this study identified a NOAEL of 75 mg/kg-day and a LOAEL of 150 mg/kg-day.
Thornton-Manning et al. (1994) conducted an analogous experiment with female
C57BL/6J mice. Six mice per group were administered an aqueous solution (10% Emulphor®) of
bromodichloromethane by gavage for five consecutive days at dose levels of 0, 75 and 150
mg/kg-day. Animals were sacrificed on day 6. All mice survived to the termination of the
experiment. No effect on body, kidney, or liver weight was observed with the exception of a
significant increase in absolute liver weight at 150 mg/kg-day. No change in cytochrome P450
activity was observed, although a nonsignificant dose-related decrease in total P450 content was
observed. ALT was significantly increased at 150 mg/kg-day, and a significant dose-related
increase in SDH activity was observed. Creatinine and BUN were not significantly increased. No
kidney or liver lesions were observed at either dose. Based on increases in serum enzyme activity,
a LOAEL of 150 mg/kg-day and a NOAEL of 75 mg/kg-day were identified for this study.
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Potter et al. (1996) investigated hyaline droplet formation and cell proliferation in the
kidney of male F344 rats. Test animals (4/dose) received 0.75 or 1.5 mmol/kg of
bromodichloromethane in 4% Emulphor® by gavage for 1, 3, or 7 days. The administered doses
corresponded to 123 or 246 mg/kg-day. No exposure-related increase in hyaline droplet
formation was observed at either dose. Binding of bromodichloromethane to a2u-globulin was not
measured. Cell proliferation in the kidney was assessed in vivo by [3H]-thymidine incorporation.
No statistically significant effect of bromodichloromethane on tubular cell proliferation was
observed following exposures of up to 7 days, although high labeling levels were observed in 3 of
4 rats at the 246 mg/kg-day dose.
Melnick et al. (1998) exposed female B6C3F, mice (10 animals/group) to
bromodichloromethane in corn oil via gavage for 3 weeks (5 days/week). Doses of
bromodichloromethane used in this study were 0 (vehicle only), 75, 150, or 326 mg/kg-day.
There were no treatment related signs of overt toxicity observed during the study. Body weight
and water intake were not significantly altered at any dose tested. However, a significant dose-
related increase in absolute liver weight and liver weight/body weight ratio was noted for the 150
and 326 mg/kg-day dose groups. Serum ALT activity was significantly increased in the two
highest dose groups and serum SDH activity was elevated at all doses tested. At necropsy, there
was clear evidence of hepatocyte hydropic degeneration in animals treated with 150 and 326
mg/kg-day. BrdU was administered to the animals during the last 6 days of the study, and
hepatocyte labeling index (LI) analysis was conducted. The two highest (150 and 326 mg/kg-
day) doses resulted in significantly elevated hepatocyte proliferation as measured by the LI.
NOAEL and LOAEL values of 75 and 150 mg/kg-day were identified on the basis of elevated
serum enzyme activity, increased liver weight, and histological findings.
NTP (1998) evaluated the effect of bromodichloromethane on food and water
consumption by Sprague-Dawley rats in the course of a range-finding experiment for a study of
developmental and reproductive effects. This study was conducted in compliance with the Good
Laboratory Practice Regulations as described in 21 CFR 58. Bromodichloromethane was
administered to test animals (6 animals/sex/dose) at nominal concentrations of 0, 100, 500, 1000,
and 1500 ppm in the drinking water for 2 weeks. The average doses of bromodichloromethane
estimated based on water consumption were 11, 45, 91 and 124 mg/kg-day for the 100, 500,
1000 and 1500 ppm dose groups, respectively. All animals were observed twice daily for signs of
toxicity. Body weight data were obtained twice weekly and at termination of the experiment.
Feed and water consumption were measured twice weekly. Animals were euthanized at
termination of the experiment without necropsy. No mortality or treatment-related clinical signs
were observed in any dose group. Body weights and weight gains were comparable among all
dose groups, except for body weight gains on Study Day 5 (the first day of compound
administration) in the 1000 and 1500 ppm dose groups which were decreased 127.5% and
118.5%, respectively. Feed consumption was also comparable across dose groups, with the
exception of male rats dosed with 1000 and 1500 ppm. Male rats in these dose groups showed
decreases in consumption of 31% and 41% , respectively, on Study Days 1 to 5. Water
consumption was reduced in the 500, 1000, and 1500 ppm dose groups, suggesting that
bromodichloromethane is unpalatable at higher concentrations. The greatest reduction in water
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intake was noted on Study Days 1 to 5 (61% and 62% for males in the 1000 ppm and 1500 ppm
dose groups, respectively, and 38%, 40% and 52% for females in the 500, 1000 and 1500 ppm
dose groups, respectively).
NTP (1998) conducted a short-term reproductive and developmental toxicity screen in
Sprague-Dawley rats to evaluate the potential toxicity of bromodichloromethane administered in
drinking water for 35 days. This study was conducted in compliance with the Good Laboratory
Practice Regulations as described in 21 CFR 58. Groups of male and female rats (5-13/sex/dose)
were exposed to drinking water concentrations of 0, 100, 700 and 1300 ppm
bromodichloromethane using the study design described in Table V-6 (Section V.E. 1). Feed and
water consumption, body weight, hematology, clinical chemistry, cell proliferation, and pathology
were evaluated in addition to developmental and reproductive endpoints. Based on water
consumption and analytical measurements of bromodichloromethane in the provided drinking
water, the calculated average daily doses were 0, 9, 38, and 67 mg/kg-day for Group A males; 0,
7, 43, and 69 mg/kg-day for Group B males; and 0, 14, 69, or 126 mg/kg-day for Group C
females.
The results for reproductive and developmental effects are reported in Section V.E. 1.
Alterations in hematological endpoints or clinical chemistry were not observed following
bromodichloromethane exposure, with the exception of a 14% drop in creatinine in the 100 ppm
Group A males and a 43% increase in 5-nucleotidase in the 1300 ppm Group A males when
compared to controls. An increase in 5-nucleotidase is an indication of hepatobiliary dysfunction
in which there is interference with the secretion of bile, and should be accompanied by a parallel
change in alkaline phosphatase activity. Since alkaline phophatase activity was unaltered in this
study, the toxicological significance of the observed increase in 5-nucleotidase was considered
uncertain. Organ weight and organ/body weight ratios reported by NTP (1998) were comparable
in all treatment groups for both males and females. Histopathological examination identified three
tissue changes that were potentially treatment-related. Cytoplasmic vacuolization of hepatocytes
and mild liver necrosis were observed in Group A males (see Table V-6 for details of group
assignment) treated with 700 and 1300 ppm bromodichloromethane and in Group B males treated
with 1300 ppm bromodichloromethane. Hepatic necrosis was dose-dependent, with incidences of
0/10, 0/10, 4/9, and 10/10 observed at 0, 100, 700, and 1300 ppm, respectively. These changes
were not accompanied by an increase in alkaline phosphatase activity. Hematopoietic cell
proliferation in the spleen was observed in Group A males at all doses of bromodichloromethane.
However, the biological significance of this finding with respect to bromodichloromethane
treatment was unclear, since cell proliferation in the spleen may occur as a response to general
stress. Evidence of mild kidney necrosis was evident in Group A males in the 1300 ppm dose
group, but may have resulted from decreased water intake. BrdU labeling index (LI), a
measurement of cell proliferation, was unchanged in the livers and kidneys of Group B males in all
dose groups. A small but statistically significant increase in the LI was noted in the livers and
kidneys of Group C females in the 1300 ppm dose group.
As discussed in Sections V.E.I, results from this study indicate that bromodichloro-
methane did not result in reproductive or developmental toxicity at drinking water concentrations
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up to 1300 ppm. However, exposure to concentrations of 700 ppm and 1300 ppm produced
changes in liver histopathology in male rats and resulted in decreases in body weight and food and
water consumption in both sexes. On the basis of these results, NTP (1998) concluded that
bromodichloromethane is unpalatable at these concentrations and is a possible general toxicant in
male and female rats at concentrations of 700 ppm and above. Although not accompanied by
changes in alkaline phosphatase activity, the occurrence of individual hepatocyte cell necrosis was
clearly dose-related and thus considered appropriate for identification of NOAEL and LOAEL
values. Based on calculated average daily doses for Group A males at the 100 and 700 ppm
concentrations (Table 6A in NTP, 1998), these data identify NOAEL and LOAEL values of 9
mg/kg-day and 38 mg/kg-day, respectively, for occurrence of hepatic cell necrosis.
Coffin et al. (2000) examined the effect of bromodichloromethane administered by gavage
in corn oil or in drinking water on cell proliferation and DNA methylation in the liver of female
B6C3F1 mice. Gavage doses of 0, 0.92, or 1.83 mmol/kg (0, 150, or 300 mg/kg, respectively)
were administered to test animals (7-8 weeks old; 10/group) daily for five days, off for two days,
and then again daily for four days. The high dose was selected on the basis that it had previously
been shown to be carcinogenic in female mice. In a separate experiment, bromodichloromethane
was administered in drinking water for 11 days at approximately 75% of the saturation level,
resulting in an average daily dose of 0.85 mmol/kg (138 mg/kg). The mice were sacrificed 24
hours after the last gavage dose and the livers were removed, weighed, and processed for
histopathological examination, proliferating cell nuclear antigen - labeling index (PCNA-LI)
analysis, and determination of c-myc methylation status. A significant, dose-dependent increase in
relative liver weight was observed in animals dosed by gavage; however, relative liver weight was
unaffected in animals administered the compound in drinking water, when compared to controls.
Histopathological findings in gavage-dosed animals consisted of hydropic degeneration at the low
dose and necrosis, fibrosis, and giant cell reaction at the high dose. No severity or incidence data
were provided. The histopathology findings for animals receiving bromodichloromethane in the
drinking water were similar to those observed in the low dose gavage group. Bromodichloro-
methane administered by gavage caused a dose-dependent increase in the PCNA-LI which was
significant at each dose tested when compared to the control. There was no significant effect
when the compound was administered in drinking water. Administration of
bromodichloromethane by gavage or in drinking water decreased methylation of the c-myc gene.
A LOAEL of 150 mg/kg, the lowest dose tested, was identified on the basis of liver toxicity
(hydropic degeneration) and increased cell proliferation in animals administered
bromodichloromethane by corn oil gavage. A NOAEL was not identified. The results of the
single-dose drinking water experiment suggest a slightly lower LOAEL of 138 mg/kg-day, based
on hydropic degeneration of the liver.
Torti et al. (2001) conducted a one week inhalation exposure study of
bromodichloromethane in male wild-type (p53+/+) and genetically engineered p53 heterozygous
(p53+/") mice. C57BL/6, FVB/N, and C57BL/6 p53+/" mice (6 mice/type/concentration) were
exposed to target exposure concentrations of 0, 1, 10, 30, 100, or 150 ppm for six hours per day,
seven days per week. FVB/N p53+/" mice were exposed to concentrations of 0, 0.3, 1, 10, or 30
ppm for six hours per day, seven days per week. The test animals were evaluated for clinical and
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pathological changes and induced regenerative cell proliferation in kidney and liver. Osmotic
pumps for delivery of bromodeoxyuridine for determination of labeling index were implanted at
3.5 days prior to scheduled termination. Test animals were euthanized approximately 18 hours
after the last scheduled exposure. Average measured concentrations of bromodichloromethane in
the middle dose range were 102 to 104% of the target concentration (coefficients of variation 2.6
to 10.6%). The lowest dose average concentration was 114% (coefficient of variation 38.6%).
The average high dose concentration was 78.8% of the target concentration as a result of
technical problems with the metering system. Deaths were observed at concentrations of 30 ppm
and greater, with 100% mortality observed for C57BL/6 heterozygous and FVB/N wild type mice
at 150 ppm. Clinical signs in mice surviving exposure at 100 and 150 ppm included lethargy and
labored breathing. Reddened skin and eyes were also observed, primarily at concentrations of 30
ppm and above. Relative body weight was decreased at exposure concentrations of 30 ppm and
above. Body weight loss was greater in the heterozygous p53+/" than in the corresponding wild
type strain. Significantly increased relative kidney weight was evident in all mice exposed to
concentrations of 30 ppm and above and in FVB/N mice exposed to 10 ppm. Significantly
increased relative liver weight was observed in wild type and heterozygous FVB/N mice exposed
to concentrations of 10 ppm and above. Reduced body weight gains were observed in surviving
mice exposed to 30 ppm and above, with greater reductions observed in heterozygous mice when
compared to wild type mice. Histopathologic evaluation revealed severe renal damage consisting
of nephrosis, tubular degeneration, and associated regeneration. The averaged severity scores
indicated greater damage in heterozygous compared to wild type mice and in FVB/N compared to
the C57BL/N mice. Centrilobular degeneration and necrosis were observed in the livers of
moribund mice sacrificed before study termination. These lesions were also observed in surviving
animals at study termination. Hydropic degeneration occurred at concentrations of 30 ppm and
above in wild type and heterozygous C57BL/N mice and at concentrations of 10 ppm and above
in FVB/N mice. Necrosis was evident at concentrations of 100 and 150 ppm. No histopathologic
lesions were observed in the bladder. Regenerative cell-proliferation in the kidney cortex was
significantly increased at exposure concentrations of 10 ppm and above. Renal labeling indices
were approximately 20 to 30 % at 10 ppm and 45 to 60% at the high dose. Regenerative cell
proliferation in the liver was less pronounced than in the kidney. Minimal increases in the labeling
index (approximately 10% or less) were observed in wild type C57BL/N and FVB/N mice.
Labeling index was significantly increased only in C57BL/N wild type mice exposed at the 100
ppm level. A modest increase in labeling index was observed in heterozygous FVB/N mice
exposed to 10 or 30 ppm; the response reached statistical significance at 30 ppm. Relatively large
increases in labeling index (up to 40%) were observed in heterozygous C57BL/N mice at doses of
10 ppm and above. The response was statistically significant at the 30 and 100 ppm exposure
levels. Bromodichloromethane did not induce cellular proliferation in the transitional epithelium
of the bladder. These data identify NOAEL and LOAEL values of 1 and 10 ppm, respectively,
based on histopathological changes in the kidney of male p53 wild type and heterozygous
C57BL/6 and FVB/N mice.
Torti et al. (2001) also conducted a three week inhalation exposure study of
bromodichloromethane in wild-type (p53+/+) and genetically engineered p53 heterozygous (p53+/")
male mice. C57BL/6, FVB/N, C57BL/6 p53+/", and FVB/N p53+/" mice (6
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mice/type/concentration) were exposed to target exposure concentrations of 0.3, 1, 3, 10, or 30
ppm for six hours per day, seven days per week. The test protocol and endpoints measured were
the same as those used for the one week study described above. Test animals were euthanized
approximately 18 hours after the last scheduled exposure. Average measured concentrations
were 92 to 97% of the target concentrations, with coefficients of variation ranging from 3.1 to
6.9%. Mortality was observed in all 30 ppm dose groups with the exception of wild type
C57BL/6 mice. No clinical signs of toxicity were reported. Body weight gain was significantly
reduced only in C57BL/6 wild type mice exposed at 30 ppm. Relative kidney weights in exposed
groups did not differ significantly from the control values. Significantly increased relative liver
weight was observed in only in heterozygous C57BL/6 and wild type FVB/N mice exposed at 30
ppm. Histopathologic evaluation revealed near-normal kidney architecture. Minimal to moderate
degenerative tubular change and regenerative tubules were observed in the 10 and 30 ppm
groups, but the acute tubular nephrosis observed in the one week study was not evident. Minimal
hepatocyte degeneration was observed in heterozygous C57BL/6 mice exposed at 30 ppm and in
heterozygous FVB/N mice exposed at 10 or 30 ppm. These observations suggest that the liver
and severe renal toxicity observed in the one week experiment conducted by Torti et al. (2001)
are transient and were resolving by three weeks. No histopathologic lesions were observed in the
bladder. Regenerative cell-proliferation in the kidney cortex was near baseline levels, with only
the 30 ppm groups showing small elevations. These elevations were statistically significant in all
30 ppm groups except C57BL/N wild type mice. No increases in regenerative cell proliferation
were evident in the liver or bladder. The NOAEL and LOAEL values in this study are 3 and 10
ppm, respectively, based on histopathologic changes in the liver and kidney of male p53 wild type
and heterozygous C57BL/6 and FVB/N mice.
2. Dibromochloromethane
Chu et al. (1982a) administered dibromochloromethane to male Sprague-Dawley rats
(10/group) in drinking water for 28 days at dose levels of 0, 5, 50, or 500 ppm. Based on
recorded fluid intake, these levels corresponded to doses of 0, 0.7, 8.5, or 68 mg/kg-day, as
calculated by the authors. The authors observed no effects on growth rate or food consumption
and no signs of toxicity throughout the exposure. No dose-related biochemical or histologic
changes were detected (no data provided). This study identified a NOAEL of 68 mg/kg-day, but
the reported data were too limited to allow an independent verification.
Munson et al. (1982) administered dibromochloromethane by aqueous gavage to male and
female CD-I mice (8 to 12/sex/group) for 14 days at dose levels of 0, 50, 125, or 250 mg/kg-day.
Endpoints measured included body and organ weights, hematology, clinical chemistry, and
humoral and cell-mediated immune system function. At 250 mg/kg-day, body weights were
significantly decreased only in high-dose males. Significant organ weight changes included
increased absolute liver weight (high-dose females), increased relative liver weight (mid- and high-
dose groups), and decreased absolute and relative spleen weight (high-dose males). The only
hematology parameter significantly affected by treatment was fibrinogen concentration, which was
decreased in high-dose males and females. Significant clinical chemistry findings were limited to
the high-dose groups. Specifically, glucose levels were significantly decreased in both males and
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females, and ALT and AST activities were significantly increased in both males and females.
Dibromochloromethane appeared to affect the humoral immune system, as judged by significantly
decreased antibody-forming cells (mid- and high-dose groups) and hemagglutination titers (high-
dose groups). The cell-mediated immune system also appeared to be affected in male animals, as
judged by a significant decrease in the popliteal lymph node stimulation index at the high dose.
This study identified a NOAEL of 50 mg/kg-day and a LOAEL of 125 mg/kg-day for
dibromochloromethane, based on decreased immune function.
Condie et al. (1983) investigated the renal and hepatic toxicity of dibromochloromethane.
Male CD-I mice (8 to 16/group) were administered 0, 37, 74, or 147 mg/kg-day of
dibromochloromethane by gavage in corn oil for 14 days. Biochemical evidence of liver damage
(elevated ALT) and kidney damage (decreased PAH uptake by kidney slices) was observed at the
high dose, but not at the mid-level or low doses. Similarly, histopathology revealed no consistent
or important changes at the low or mid doses with minimal to moderate liver and kidney injury at
the high dose. Liver lesions included mitotic figures, focal inflammation, and cytoplasmic
vacuolation, while kidney lesions included epithelial hyperplasia and mesangial nephrosis. On this
basis, this study identified a NOAEL value of 74 mg/kg-day and LOAEL value of 147 mg/kg-day
for dibromochloromethane.
In a 14-day study by NTP (1985), groups of male and female F344/N rats (5/sex/dose)
were administered 0, 60, 125, 250, 500, or 1,000 mg/kg-day of dibromochloromethane by gavage
in corn oil. Animals were observed twice daily for mortality and were weighed once per week.
Necropsies were performed on all animals. All high-dose rats and all females that received 500
mg/kg-day died by day 6. Three males at 500 mg/kg-day died between days 5 and 8. No deaths
occurred at or below 250 mg/kg-day. At 500 or 1000 mg/kg-day, clinical observations included
lethargy, ataxia, and labored breathing. Treatment-related macroscopic findings included mottled
livers and darkened renal medullae in animals administered 500 or 1,000 mg/kg-day. Based on
behavior, gross pathology, and mortality, this study identified a NOAEL of 250 mg/kg-day and a
LOAEL of 500 mg/kg-day.
In a parallel study (NTP, 1985), male and female B6C3F, mice (5/sex/dose) were
administered 0, 30, 60, 125, 250, or 500 mg/kg-day of dibromochloromethane in corn oil by
gavage for 14 days. Treatment-related deaths occurred in 80% of the males and in 60% of the
females at the high dose. Clinical signs at this dose included lethargy, ataxia, and labored
breathing. Treatment-related macroscopic findings included mottled livers and darkened renal
medullae in high-dose males and females. White papillomatous nodules in the stomach were also
observed in males at 125, 250, or 500 mg/kg-day and in female mice at 250 or 500 mg/kg-day.
Based on gross lesions, this study identified a NOAEL of 60 mg/kg-day and a LOAEL of
125 mg/kg-day in mice.
Aida et al. (1992a) investigated the effects of administering dibromochloromethane to
Slc:Wistar rats (7/sex/group) for one month at dietary levels of 0%, 0.020%, 0.062%, or 0.185%)
for males, and 0%>, 0.038%>, 0.113%>, or 0.338%> for females. The test material was
microencapsulated and mixed with powdered feed; placebo granules were used for the control
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groups. Based on the mean food intakes, the study authors reported calculated doses of 0, 18.3,
56.2, or 173.3 mg/kg-day for males and 0, 34.0, 101.1, or 332.5 mg/kg-day for females. Clinical
effects, body weight, food consumption, hematology parameters, serum chemistry, and
histopathology of all major organs were determined. Body weights were significantly reduced in
high-dose females relative to the controls. High-dose females also exhibited slight piloerection
and emaciation. Dose-related increases in both absolute and relative liver weights were observed
in males (significant at the high dose) and females (significant at all dose levels with the exception
of a nonsignificant increase in absolute liver weight at the low dose). Relative kidney weights
were also significantly increased in the high-dose females. Significant decreases in alkaline
phosphatase (mid- and high-dose males and all female dose groups) and LDH (all female dose
groups) were observed, but the biological significance of these changes is unclear. Significant,
dose-related changes in serum biochemistry included reduced nonesterified fatty acids in high-
dose males, reduced T-GLY in high-dose groups, and increased cholesterol in mid- and high-dose
males and in females at all dose levels. The cholesterol levels, however, were within normal
ranges at all dose levels. Serum cholinesterase activity was also significantly decreased in high-
dose males and mid- and high-dose females with the trend clearly dose-related in females. Liver
cell vacuolization was generally noted at a similar incidence in the controls and all dosing groups,
but dose-related increases in severity were observed in mid- and high-dose males and females.
The incidence and very slight severity of the effects at the low dose were similar to those
observed in the control groups and were not considered adverse. The severity of the liver cell
vacuolization at the mid-dose was rated as very slight to slight, while the severity at the high-dose
was rated as moderate to remarkable. Swelling and single cell necrosis were also observed,
primarily in the high-dose groups. No effect was observed on any hematology parameter. Based
on the histopathology findings, NOAELs of 18.3 (males) and 34.0 (females) mg/kg-day and
LOAELs of 56.2 (males) and 101.1 (females) mg/kg-day were identified for
dibromochloromethane in rats.
Potter et al. (1996) evaluated hyaline droplet formation and cell proliferation in the kidney
of male F344 rats following exposure to dibromochloromethane. The rats (4/dose)were dosed
with 0.75 or 1.5 mmol/kg (156 or 312 mg/kg-day, respectively) of dibromochloromethane in 4%
Emulphor® by gavage for 1, 3, or 7 days. No exposure-related increase in hyaline droplets was
observed in dosed rats. Binding to a2u-globulin was not measured. Changes in kidney tubule cell
proliferation were assessed by in vivo incorporation of [3H]-thymidine. No statistically significant
effect of dibromochloromethane exposure on this endpoint was noted following exposures of up
to 7 days duration.
Melnick et al. (1998) exposed female B6C3F, mice (10/dose) to dibromochloromethane in
corn oil via gavage for 3 weeks (5 days/week). The doses of dibromochloromethane in this study
were 0 (vehicle only), 50, 100, 192, or 417 mg/kg-day. The corresponding time-weighted doses
were 0, 37, 71, 137, and 298 mg/kg-day. No treatment-related signs of overt toxicity were
observed during the study. Body weight and water intake were not significantly altered at any
dose tested. However, a statistically significant and dose-related increase in liver weight/body
weight ratio was seen in the 100, 192 and 417 mg/kg-day dose groups. Serum ALT activity was
significantly increased in the two highest dose groups. The activity of serum SDH was
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significantly elevated at all doses tested except 50 mg/kg-day. However, the increase in activity
(shown graphically) was very small relative to the control at the 100 and 192 mg/kg-day doses.
At necropsy, there was clear evidence of hepatocyte hydropic degeneration in the 192 and 417
mg/kg-day dose groups. BrdU was administered to the animals during the last 6 days of the
study, and hepatocyte labeling index (LI) analysis was conducted. Only the highest dose tested
(417 mg/kg-day) resulted in significantly elevated hepatocyte proliferation as measured by the LI.
Evaluation of the data in this study suggest a LOAEL of 192 mg/kg-day, based on a consistent
pattern of positive results for indicators of hepatotoxicity at this dose.
Coffin et al. (2000) examined the effect of dibromochloromethane administered by gavage
in corn oil or in drinking water on cell proliferation and DNA methylation in the liver of female
B6C3F1 mice. Gavage doses of 0, 0.48, or 1.44 mmol/kg (0, 100, or 300 mg/kg, respectively)
were administered to test animals (7-8 weeks old; 10/group) daily for five days, off for two days,
and then again daily for four days. The high dose was selected on the basis that it had previously
been demonstrated to be carcinogenic in female mice. Dibromochloromethane was administered
in drinking water at approximately 75% of the saturation level, resulting in an average daily dose
of 0.82 mmol/kg (171 mg/kg). The mice were sacrificed 24 hours after the last gavage dose and
the livers were removed, weighed, and processed for histopathological examination, proliferating
cell nuclear antigen - labeling index (PCNA-LI) analysis, and determination of c-myc methylation
status. For histopathological analysis, stained liver sections were evaluated for toxicity using a
semi-quantitative procedure using the following severity scoring system: Grade 1 consisted of mid
lobular ballooning hepatocytes; Grade 2 consisted of mid lobular ballooning hepatocytes
extending to the central vein; Grade 3 consisted of centrilobular necrosis with ballooning
hepatocytes; and Grade 4 consisted of necrosis extending from the central vein to the mid lobule
zone. A significant, dose-dependent increase in relative liver weight was observed in animals
dosed by gavage; however, relative liver weight was unaffected in animals administered the
compound in drinking water, when compared to controls. At the low gavage dose, liver toxicity
consisted mainly of a Grade 1 response. At the high dose, liver toxicity consisted mainly of a
Grade 2 response. No incidence data were provided in the study report, nor was a severity grade
reported for the control group. The histopathology findings for animals receiving
bromodichloromethane in the drinking water were similar to those observed in the low dose
gavage group. Dibromochloromethane administered by gavage caused a dose-dependent
increase in the PCNA-LI. The increases observes at each dose were significantly different from
the control. There was no significant effect on PCNA-LI when the compound was administered
in drinking water. Administration of dibromochloromethane by gavage or in drinking water
decreased methylation of the c-myc gene. A LOAEL of 100 mg/kg, the lowest dose tested, was
identified on the basis of liver toxicity (ballooning hepatocytes) and increased cell proliferation in
gavaged animals.
3. Bromoform
Chu et al. (1982a) administered bromoform to male Sprague-Dawley rats (10/group) in
drinking water for 28 days at dose levels of 0, 5, 50, or 500 ppm. Based on recorded fluid intake,
these levels corresponded to doses of 0, 0.7, 8.5, or 80 mg/kg-day, as calculated by the authors.
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The authors observed no effects on growth rate or food consumption and no signs of toxicity
throughout the exposure. No dose-related biochemical or histologic changes were detected (no
data provided). This study identified a NOAEL of 80 mg/kg-day, but the reported data were too
limited to allow an independent confirmation.
Munson et al. (1982) administered bromoform by aqueous gavage to male and female CD-
1 mice (6 to 12/sex/group) for 14 days at levels of 0, 50, 125, or 250 mg/kg-day. Parameters
observed included body and organ weights, hematology, clinical chemistry, and humoral and cell-
mediated immune system functions. Body weights were significantly decreased in high-dose
females, while body weights in males were significantly increased at the mid and high doses.
Absolute and relative liver weights were significantly increased in males at the mid and high dose
and in females at the high dose. Absolute spleen weight was also decreased in mid- and high-dose
females. Hematologic effects included significantly decreased fibrinogen in males at the high dose
and significantly decreased prothrombin time in all treated males. The changes in prothrombin
time, however, were not dose-related. Significant clinical chemistry findings included decreased
glucose levels (high-dose males), increased AST activity (high-dose groups), and decreased BUN
levels (high-dose males). Both the humoral and cell-mediated immune systems appeared to be
affected in males at the high dose with a significant decrease in antibody-forming cells and a
significant decrease in delayed-type hypersensitivity response. The authors stated that no
treatment-related effects on the immune system in females were observed (no data were
reported). Based on changes in clinical chemistry parameters, this study identified a NOAEL of
125 mg/kg-day and a LOAEL of 250 mg/kg-day.
Condie et al. (1983) investigated the renal and hepatic toxicity of bromoform. Male CD-I
mice (8 to 16/group) were administered 0, 72, 145, or 289 mg/kg-day of bromoform by gavage in
corn oil for 14 days. Biochemical evidence of liver damage (elevated ALT) and kidney damage
(decreased PAH uptake by kidney slices) was observed at the high dose, but not at the mid or low
dose. Histopathological examination revealed no consistent or important changes at the low or
mid doses, with minimal to moderate liver and kidney injury at the high dose. Specific
microscopic changes included intratubular mineralization, epithelial hyperplasia, mesangial
hypertrophy and mesangial nephrosis in the kidney, and centrilobular pallor, mitotic figures, focal
inflammation, and cytoplasmic vacuolation in the liver. On this basis, this study identified a
NOAEL value of 145 mg/kg-day and a LOAEL value of 289 mg/kg-day.
NTP (1989a) investigated the short term oral toxicity of bromoform in F344/N rats and
B6C3F, mice. Groups of male and female rats (5/sex/group) and female mice (5/group) were
administered doses of 0, 100, 200, 400, 600, or 800 mg/kg-day of bromoform in corn oil by
gavage for 14 days. Male mice were administered 0, 50, 100, 200, 400, or 600 mg/kg-day. All
rats that were dosed at 600 or 800 mg/kg-day died before the end of the study. At 400 mg/kg-
day, only one male rat died before study termination. These rats exhibited lethargy, labored
breathing, and ataxia. At 400 mg/kg-day, final body weights were decreased by 14% in male rats
and by 4% in female rats relative to controls. In mice, one male and one female administered the
high dose died before study termination. At dose levels of 600 mg/kg-day or above, ataxia and
lethargy were noted. Final body weights of mice were comparable to those of the controls.
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Raised stomach nodules were observed in males at 400 and 600 mg/kg-day and in females at
600 and 800 mg/kg-day. Based on decreased body weight and mortality in rats and on stomach
nodules in mice, this study identified a NOAEL of 200 mg/kg-day and a LOAEL of 400 mg/kg-
day.
Aida et al. (1992a) administered bromoform to Slc:Wistar rats (7/sex/group) for one
month at dietary levels of 0%, 0.068%, 0.204%, or 0.612%) for males and 0%, 0.072%, 0.217%),
or 0.651%> for females. The test material was microencapsulated and mixed with powdered feed;
placebo granules were used for the control groups. Based on the mean food intakes, the study
authors reported the mean compound intakes as 0, 61.9, 187.2, or 617.9 mg/kg-day for males and
0, 56.4, 207.5, or 728.3 mg/kg-day for females. Clinical effects, body weight, food consumption,
hematology parameters, serum chemistry, and histopathology of all major organs were
determined. Body weights were significantly reduced in high-dose males relative to the controls.
High-dose animals of both sexes exhibited slight piloerection and emaciation. Relative liver
weight was significantly increased in mid- and high-dose males and females. Significant changes
in serum chemistry were primarily observed in the mid- and high-dose animals with the females
more significantly affected. These changes included significant decreases in (a) serum glucose in
low- and high-dose males and in mid- and high-dose females, (b) triglycerides in high-dose males
and in mid- and high-dose females, (c) cholinesterase activity in high-dose males and in all female
treatment groups, (d) LDH in mid- and high-dose females, and (e) BUN in mid- and high-dose
females. All of these changes in the groups noted exhibited strong dose-related trends with the
exception of serum glucose in males. Creatinine levels and alkaline phosphatase activity were also
significantly decreased in all female treatment groups, but the changes were not dose-related.
Significant increases, although not dose-related, were observed for phospholipids and cholesterol
in mid- and high-dose animals with the exception of a nonsignificant increase in phospholipids in
high-dose females. The only change of clear biological significance at the low dose was a
decrease in cholinesterase activity in females. No effect was observed on any hematology
parameter. Microscopic and macroscopic findings were limited to the liver. Specifically,
discoloration was observed in all males and females in the high-dose group. The incidence and
severity of liver cell vacuolization and swelling were dose-related. Severe hepatic cell
vacuolization was observed in 5/7 high-dose males and in 6/7 females at the mid and high dose.
Slight to moderate liver cell swelling was observed in three high-dose males, while all high-dose
females displayed slight signs of liver cell swelling. Females appeared to be more sensitive for
development of histopathological effects, but the changes observed in low-dose females were not
considered an adverse effect. Based on the histopathology and serum chemistry changes in the
mid-dose animals, this study identified NOAELs of 61.9 mg/kg-day for males and 56.4 mg/kg-day
for females, and LOAELs of 187.2 mg/kg-day for males and 207.5 mg/kg-day for females.
Potter et al. (1996) evaluated the effect of bromoform on hyaline droplet formation and
cell proliferation in the kidney of male F344 rats. Animals (4/dose) received doses of 0.75 or 1.5
mmol/kg of bromoform in 4% Emulphor® by gavage for 1, 3, or 7 days. These doses correspond
to 190 or 379 mg/kg-day, respectively. No exposure-related increase in hyaline droplet
formation was observed. Cell proliferation in the kidney following bromoform exposure was
measured in vivo by [3H]-thymidine incorporation. No statistically significant effects were noted
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following exposures of up to 7 days duration. Binding of bromoform to a2u-globulin was not
measured.
Melnick et al. (1998) exposed female B6C3F, mice (10 animals/group) to bromoform in
corn oil via gavage for 3 weeks (5 days/ week). Doses of bromoform used in this study were 0
(vehicle only), 200, or 500 mg/kg-day. There were no treatment-related signs of overt toxicity
observed during the study. Body weight and water intake were not significantly altered at any
dose tested. However, a dose-related increase in absolute liver weight and liver weight/body
weight ratio was noted in both tested doses. Neither serum ALT nor serum SDH activity were
significantly elevated at either dose of bromoform. At necropsy, there was no evidence of
hepatocyte hydropic degeneration in animals treated with either dose. BrdU was administered to
the animals during the last 6 days of the study, and hepatocyte labeling index (LI) analysis was
conducted. Only the 500 mg/kg-day dose resulted in marginally significant increase in hepatocyte
proliferation as measured by the LI. These data suggest a NOAEL of 200 mg/kg-day and a
LOAEL of 500 mg/kg-day based on increased hepatocyte proliferation.
Coffin et al. (2000) examined the effect of bromoform administered by gavage in corn oil
or in drinking water on liver toxicity, cell proliferation and DNA methylation in female B6C3F1
mice. Gavage doses of 0, 0.79, or 1.98 mmol/kg (0, 200, or 500 mg/kg, respectively) were
administered to test animals (7-8 weeks old; 10/group) daily for five days, off for two days, and
then again daily for 5 days. The high dose was selected on the basis that it had previously been
demonstrated to be carcinogenic in female mice. Bromoform was administered in drinking water
at approximately 75% of the saturation level, resulting in an average daily dose of 1.19 mmol/kg
(301 mg/kg). The mice were sacrificed 24 hours after the last gavage dose and the livers were
removed, weighed, and processed for histopathological examination, proliferating cell nuclear
antigen - labeling index (PCNA-LI) analysis, and determination of c-myc methylation status. For
histopathological analysis, stained liver sections were evaluated for toxicity using a semi-
quantitative procedure using the following severity scoring system: Grade 1 consisted of mid
lobular ballooning hepatocytes; Grade 2 consisted of mid lobular ballooning hepatocytes
extending to the central vein; Grade 3 consisted of centrilobular necrosis with ballooning
hepatocytes; and Grade 4 consisted of necrosis extending from the central vein to the mid lobule
zone. A significant, dose-dependent increase in relative liver weight was observed in animals
dosed by gavage; however, relative liver weight was unaffected in animals administered the
compound in drinking water, when compared to controls. At the low gavage dose, liver toxicity
consisted mainly of a Grade 1 response. At the high dose, liver toxicity consisted mainly of a
Grade 2 response. No incidence data were provided in the study report, nor were severity data
presented for the control group. The histopathology findings for animals receiving bromoform in
the drinking water were similar to those observed in the low dose gavage group. Bromoform
administered by gavage caused a significant, dose-dependent increase in the PCNA-LI.
Bromoform also significantly enhanced cell proliferation when the compound was administered in
drinking water. Administration of bromoform by gavage or in drinking water decreased
methylation of the c-myc gene. A LOAEL of 200 mg/kg, the lowest dose tested, was identified
on the basis of liver toxicity in gavaged animals.
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C. Subchronic Exposure
This section addresses studies of brominated trihalomethanes that are of approximately 90
days in duration. Table V-4 summarizes the details of these subchronic studies.
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Table V-4 Summary of Subchronic Toxicity Studies for Brominated Trihalomethanes
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromodichloromethane
Oral Exposure
Chu et al.
(1982b)
Rat
SD*
Drinking
water
M
20
90 days
0
0.57
6.5
53
212
Non dose-dependent hepatic
and thyroid lesions
Chu et al.
(1982b)
Rat
SD
Drinking
water
F
20
90 days
0
0.75
6.9
57
219
Non dose-dependent hepatic
and thyroid lesions
NTP (1987)
Rat
F344/N
Gavage
(corn oil)
M,
F
10
13 weeks
(5 d/wk)
0
19
38
75 (NOAEL)
150 (LOAEL)
300
Reduced body weight gain
NTP (1987)
Mouse
B6C3F;
Gavage
(corn oil)
M
10
13 weeks
(5 d/wk)
0
6.3
13
25
50 (NOAEL)
100 (LOAEL)
Focal necrosis of proximal
renal tubular epithelium
NTP (1987)
Mouse
B6C3F;
Gavage
(corn oil)
F
10
13 weeks
(5 d/wk)
0
25
50
100 (NOAEL)
200 (LOAEL)
400
Hepatic microgranulomas
Inhalation Exposure
Torti et al.
2001
Mouse
C57BL/6
FVB/N
p53
(hetero-
zygous)
Vapor
M
Not
reported
13 weeks
(6 h/day)
0 ppm
0.5 ppm
3 ppm
10 ppm
15 ppm
Text reported minimal cortical
scarring and occasional
regenerative tubules in
C56BL/6 mice and mild renal
cortical tubular
karyocytomegaly.
Concentrations at which these
effects occurred were not
reported.
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Table V-4 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Dibromochloromethane
Chu et al.
(1982b)
Rat
SD
Drinking
water
M
20
90 days
0
0.57
6.1
49 (NOAEL)
224 (LOAEL)
Hepatic lesions
Chu et al.
(1982b)
Rat
SD
Drinking
water
F
20
90 days
0
0.64
6.9
55 (NOAEL)
236 (LOAEL)
Hepatic lesions
NTP (1985)
Rat
F344/N
Gavage
(corn oil)
M,
F
10
13 weeks
(5 d/wk)
0
15
30 (NOAEL)
60 (LOAEL)
125
250
Hepatic vacuolization
indicative of fatty
metamorphosis (males)
NTP (1985)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
10
13 weeks
(5 d/wk)
0
15
30
60
125 (NOAEL)
250 (LOAEL)
Fatty liver and toxic
nephropathy in males
Daniel et al.
(1990)
Rat
SD
Gavage
(corn oil)
M,
F
10
90 days
0
50 (LOAEL)
100
200
Hepatic vacuolization (males);
renal lesions (females)
Bromoform
Chu et al.
(1982b)
Rat
SD
Drinking
water
M
20
90 days
0
0.65
6.1
57 (NOAEL)
218 (LOAEL)
Hepatic lesions and vacuolation
Chu et al.
(1982b)
Rat
SD
Drinking
water
F
20
90 days
0
0.64
6.9
55 (NOAEL)
283 (LOAEL)
Hepatic lesions and vacuolation
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Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
NTP (1989a)
Rat
F344/N
Gavage
(corn oil)
M,
F
10
13 weeks
(5 d/wk)
0
12
25 (NOAEL)
50 (LOAEL)
100
200
Hepatic vacuolation in males
NTP (1989a)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
10
13 weeks
(5 d/wk)
0
25
50
100 (NOAEL)
200 (LOAEL)
400
Hepatic vacuolation in males
* SD, Sprague-Dawley
1. Bromodichloromethane
Chu et al. (1982b) administered bromodichloromethane to male and female weanling
Sprague-Dawley rats (20/sex/dose) in drinking water at levels of 0, 5, 50, 500, or 2,500 ppm for
90 days. Half of each group (10/sex/dose) was sacrificed at the end of the exposure period, and
the remaining animals were given tap water for another 90 days. As calculated by the authors
(using data on water consumption and the average initial and final body weights in the vehicle
controls and the high dose groups), these levels corresponded to doses of approximately 0, 0.57,
6.5, 53, and 212 mg/kg-day for males and 0, 0.75, 6.9, 57, and 219 mg/kg-day for females. At
2,500 ppm, food consumption was significantly depressed and significant growth suppression
occurred in both males and females. Mild histologic changes were observed in the liver and
thyroid of the male animals. Neither incidence nor severity were clearly dose-related.
Specifically, the incidence of hepatic lesions was increased in males at concentrations equal to or
greater than 50 ppm, with similar statistically significant increases in the severity of these lesions
in these dose groups compared to the control. The author noted that the hepatic lesions were
mild and similar to the control following the 90-day recovery period. Increased incidence of
thyroid lesions was also observed in males at concentrations equal to or greater than 50 ppm. The
severity of these lesions was similar to that observed in the control group. These lesions were
also mild and similar in nature to those of the control after the 90-day recovery period. The
incidence of hepatic lesions in the female treatment groups (3-5/10) was slightly increased
compared to that of the control group (0/10) with the severity significantly increased in the 50 and
2,500 ppm treatment groups, but not in the 500 ppm group. No significant numbers of females
were reported as having thyroid lesions. Lack of a clear dose-response relationship for either
incidence or severity of lesions prevented identification of reliable NOAEL or LOAEL values.
NTP (1987) administered doses of 0, 19, 38, 75, 150, or 300 mg/kg-day of
bromodichloromethane to male and female F344/N rats (10/sex/dose) by gavage in corn oil for
5 days/week for 13 weeks. The low-dose group was administered 1.9 mg/kg-day for the first
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3 weeks of the study. A necropsy was performed on all animals. Before study termination, 50%
of the males and 20% of the females in the high-dose group died. Although food consumption
was not recorded, animals in the high-dose groups appeared to eat less food. These animals were
also emaciated. At 300 mg/kg-day, final body weights of the males and females were decreased
by 55%) and 32%, respectively, relative to the controls. At 150 mg/kg-day, final body weights of
the males and females were decreased by 30% and 12%, respectively, relative to the controls.
Treatment-related lesions were observed only at the high dose. At 300 mg/kg-day in males,
centrilobular degeneration of the liver and occasional necrotic cells were observed in 4/9 animals.
Mild bile duct hyperplasia was also observed in these animals. Kidney lesions in high-dose males
consisted of degeneration of renal proximal tubular epithelial cells (4/9) and definite foci of
coagulative necrosis of the tubular epithelium (2/9). High-dose males (4/9) also exhibited
lymphoid degeneration of the thymus, spleen, and lymph nodes, and mild to moderate atrophy of
the seminal vesicles and/or prostate. Enlarged hepatocytes were observed in females (2/9) at
300 mg/kg-day. Although degeneration of the spleen, thymus, and lymph nodes was noted in
high-dose females, the extent of the atrophy was much less than that observed in males. This
study identified a NOAEL of 75 mg/kg-day and a LOAEL of 150 mg/kg-day based on reduced
body weight gain.
In a parallel experiment, NTP (1987) administered bromodichloromethane in corn oil by
gavage to male and female B6C3F, mice (10/sex/dose) for 5 days/week for 13 weeks. Doses
were 0, 6.25, 12.5, 25, 50, or 100 mg/kg-day for males and 0, 25, 50, 100, 200, or 400 mg/kg-
day for females. All animals survived to the end of the study. The final body weights of high-
dose males were decreased by 9% relative to the controls. The final body weights of females that
received 200 and 400 mg/kg-day were decreased 5% and 6%, respectively, relative to the
controls. No treatment-related clinical signs were noted. Treatment-related lesions were
observed only at 100 mg/kg-day in males and at 200 and 400 mg/kg-day in females. Kidney
lesions in high-dose males included focal necrosis of the proximal renal tubular epithelium (6/10)
and nephrosis of minimal severity (2/10). Microgranulomas were observed in the liver of 70% of
the females that received the 200 mg/kg-day dose. NOAEL and LOAEL values for female mice
were 100 and 200 mg/kg-day, respectively, based on occurrence of microgranulomas. This study
identified a NOAEL of 50 mg/kg-day and a LOAEL of 100 mg/kg-day for male mice on the basis
of liver histopathology.
Torti et al. (2001) reported results from a 13-week interim sacrifice conducted as part of
an inhalation cancer bioassay in p53 heterozygous C57BL/6 and FVB/N male mice. Test animals
were exposed to vapor concentrations of 0, 0.5, 3, 10, or 15 ppm, 6 hours/day for 13 weeks. No
exposure-related effects were noted for mortality, morbidity, relative body weight, relative kidney
or liver weight, or cell proliferation in liver, kidney or bladder. Histopathologic lesions were
limited to the kidney. The study authors reported minimal cortical scarring and occasional
regenerative tubules in the C57BL/6 strain. The only lesion reported for the FVB/N strain was
limited to mild renal cortical tubular karyocytomegaly. No incidence data were presented for
these lesions and the concentrations at which they occurred was not stated.
2. Dibromochloromethane
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Chu et al. (1982b) administered dibromochloromethane to male and female weanling
Sprague-Dawley rats (20/sex/dose) in drinking water at levels of 0, 5, 50, 500, or 2,500 ppm for
90 days. Half of each group (10/sex/dose) was sacrificed at the end of the exposure period, and
the remaining animals were given tap water for another 90 days. Based on calculations by the
authors, these levels corresponded to doses of approximately 0, 0.57, 6.1, 49, and 224 mg/kg-day
for males and 0, 0.64, 6.9, 55, and 236 mg/kg-day for females. At 2,500 ppm, food consumption
was depressed in both males and females, with the decrease reaching statistical significance in the
males. Body weight gain was also decreased at the high-dose, but not significantly. Mild
histologic changes occurred in the liver and thyroid in both males and females. Neither the
incidence nor severity exhibited clear dose-response trends with the possible exception of the
incidence and severity of hepatic lesions in the males. The severity of hepatic lesions was
significantly increased at 50 ppm in females and at 2,500 ppm in both males and females. Hepatic
lesions included increased cytoplasmic volume and vacuolation due to fatty infiltration. Lesions
of the thyroid included decreased follicular size and colloid density and occasional focal collapse
of follicles. The severity of these lesions was not significantly different from that of the control.
The authors noted that histological changes were mild and similar to controls when evaluated
after the 90-day recovery period. These data identified a NOAEL of 49 mg/kg-day and a LOAEL
of 224 mg/kg-day for males, and a NOAEL of 55 mg/kg-day and a LOAEL of 236 mg/kg-day for
females.
NTP (1985) administered dibromochloromethane by gavage in corn oil to male and female
F344/N rats (10/dose/sex). Doses of 0, 15, 30, 60, 125, or 250 mg/kg were given 5 days/week
for 13 weeks. Animals were weighed weekly. All animals were submitted for gross necropsy,
while histopathology was conducted on animals in the control and high-dose groups with the
exception that the liver was examined in all males and in females at 125 mg/kg-day, and that the
kidney and salivary glands were examined in males and females at 125 mg/kg-day. Only one male
and one female in the high-dose group survived, with most deaths occurring during weeks 8 to
10. At 125 mg/kg-day, final body weights of males were decreased 7% relative to controls. His-
topathological examination revealed severe lesions and necrosis in kidney, liver, and salivary
glands, primarily at the high-dose. Males, however, exhibited a dose-dependent increase in the
frequency of clear cytoplasmic vacuoles indicative of fatty metamorphosis in the liver; this effect
was statistically significant at doses of 60 mg/kg-day or higher. On this basis, this study identified
a NOAEL of 30 mg/kg-day and a LOAEL of 60 mg/kg-day in rats for dibromochloromethane.
BMD modeling was conducted on the incidence of fatty metamorphosis in male rats.
NTP (1985) performed a similar 13-week gavage study with dibromochloromethane in
male and female B6C3F, mice (10/sex/dose). The doses and dosing schedule were the same as
for the rat study. No treatment-related effects on body weight or histopathology were observed
at doses of 125 mg/kg-day or lower. At the high dose, final body weights of males and females
were decreased by 6% relative to controls. Fatty metamorphosis of the liver and toxic
nephropathy were observed in high-dose males, but not in high-dose females. This study
identified a NOAEL of 125 mg/kg-day and a LOAEL of 250 mg/kg-day in mice.
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Daniel et al. (1990) administered gavage doses (in corn oil) of 0, 50, 100, or 200 mg/kg-
day of dibromochloromethane to male and female Sprague-Dawley rats (10/sex/dose) for 90
consecutive days. Individual dosages were adjusted weekly based on individual body weights.
During the final week of the study, urinalysis was conducted following an overnight fast.
Ophthalmoscopic examinations were performed prior to treatment and during the last week of the
study. Hematology, serum clinical chemistry, and a thorough histopathologic examination were
also conducted. No deaths, clinical signs of toxicity, or treatment-related changes in the
ophthalmoscopic examinations or hematology were observed. Final body weights were
significantly reduced in the high-dose groups by 32% in males and by 13% in females. Body
weight decreases in the other groups were less than 10% of control weights. A dose-related
increase was observed in liver weight in females that reached statistical significance at the high
dose. Clinical chemistry values indicative of hepatotoxicity and suggestive of nephrotoxicity
included increased levels of alkaline phosphatase (high-dose males and females), ALT (mid- and
high-dose males), and creatinine (mid- and high-dose males and high-dose females), and
decreased potassium levels (high-dose males). Centrilobular lipidosis (vacuolization) was
observed in the liver of almost all high-dose males and females and all mid- and low-dose males
(with one exception at each level), but in only one mid-dose female. The severity of the effect
was dose-related. Centrilobular necrosis was also observed in high-dose males and females.
Slight-to-moderate degeneration within the kidney proximal tubular cells occurred in all high-dose
males and females and to a lesser extent in mid-dose males and low- and mid-dose females. Based
on the liver histopathology in males and kidney histopathology in females, the LOAEL for
dibromochloromethane in this study was 50 mg/kg-day.
3. Bromoform
Chu et al. (1982b) administered bromoform to male and female weanling Sprague-Dawley
rats (20 rats/sex/group) for 90 days in drinking water at levels of 0, 5, 50, 500, or 2,500 ppm.
Half of each group (10/sex/dose) was sacrificed at the end of the exposure period, and the
remaining animals were given tap water for another 90 days. Based on calculations by the
authors, these levels corresponded to doses of approximately 0, 0.65, 6.1, 57, and 218 mg/kg-day
for males and 0, 0.64, 6.9, 55, and 283 mg/kg-day for females. At 2,500 ppm, food consumption
was depressed in both males and females, with the decrease reaching statistical significance in
males. Body weight gain was also decreased at the high-dose, but not significantly. Lymphocyte
counts were significantly decreased in high-dose males and females when evaluated 90-days after
cessation of treatment. The only change in serum biochemistry was a significant decrease in LDH
in both males and females at the high dose. This effect was also noted 90-days after cessation of
treatment. Mild histologic changes occurred in the liver and thyroid of male and female animals.
Although neither incidence nor severity were clearly dose-related, these parameters did tend to
increase with dose. The severity of hepatic lesions was significantly increased in high-dose males
and in females at 500 and 2,500 ppm. Hepatic lesions included increased cytoplasmic volume and
vacuolation due to fatty infiltration. Lesions of the thyroid included decreased follicular size and
colloid density and occasional focal collapse of follicles. The severity of these lesions in the
treated animals was not significantly different from that in the controls. Although the authors
noted that histologic changes were mild and similar to controls when evaluated after the 90-day
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recovery period, males in the high-dose group continued to exhibit an increased incidence of
hepatic lesions with greater severity relative to the control. These data identified a NOAEL of
57 mg/kg-day and a LOAEL of 218 mg/kg-day for males, and a NOAEL of 55 mg/kg-day and a
LOAEL of 283 mg/kg-day for females.
NTP (1989a) exposed male and female F344/N rats to bromoform by gavage for
5 days/week for 13 weeks. Animals (10/sex/dose) received doses of 0, 12, 25, 50, 100, or
200 mg/kg-day. None of the rats died before the end of the study, and body weights were not
significantly affected. All high-dose animals, as well as males dosed with 100 mg/kg-day, were
lethargic. At sacrifice, tissues were examined for gross and histologic changes. A dose-
dependent increase in the frequency of hepatocellular vacuolation was observed in male rats,
which reached statistical significance at 50 mg/kg-day (IRIS, 1993b). These hepatic effects were
not observed in females. This study identified a NOAEL of 25 mg/kg-day and a LOAEL of
50 mg/kg-day, on the basis of the hepatic vacuolation seen in male rats.
In a parallel study, NTP (1989a) exposed male and female B6C3F, mice to bromoform by
gavage for 5 days/week for 13 weeks. Animals (10/sex/dose) received doses of 0, 25, 50, 100,
200, or 400 mg/kg-day. One female died at 100 mg/kg-day, but no other deaths at any other dose
level occurred. At sacrifice, tissues were examined for gross and histologic changes. Body
weights were not significantly affected, although males receiving 400 mg/kg-day had body
weights about 8% less than controls. A dose-related increase in the number of hepatocellular
vacuoles was seen in male mice (incidence of 5/10 at 200 mg/kg and 8/10 at 400 mg/kg reported
in text; incidence in controls not explicitly stated), but not in females. Based on hepatocellular
vacuolation, this study identified a NOAEL of 100 mg/kg-day and a LOAEL of 200 mg/kg-day in
male mice.
D. Chronic Exposure
This section addresses studies on the health effects of brominated trihalomethanes that are
of one to two years in duration. These studies are summarized in Table V-5.
1. Bromodichloromethane
Tobe et al. (1982) evaluated the chronic effects of bromodichloromethane administered in
the diet to male and female Slc:Wistar SPF rats (40/sex/group) for 24 months. The
histopathology data for the animals exposed to bromodichloromethane in this study were reported
by Aida et al. (1992b). The animals were 5 weeks old at the start of the study and weighed
approximately 100 g. Bromodichloromethane was microencapsulated, and an appropriate amount
was mixed with powdered feed. The concentrations administered were 0.0. 0.014, 0.055, or
0.22%. Control groups (70 rats/sex) received microcapsules without the test compound. Body
weight and food consumption were monitored weekly for the first 6 months, every 2 weeks from
6 to 12 months, and every 4 weeks during the second year of the study. Interim sacrifices of at
least 9 animals/sex/control group and 5 animals/sex/dose group were conducted at 6, 12, and 18
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months. All surviving animals were sacrificed at 2 years. At each time of sacrifice, necropsies,
hematology, and serum biochemistry were conducted. Based on mean food intakes, the reported
average doses were approximately 0, 6, 26, or 138 mg/kg-day for males and 0, 8, 32, or
168 mg/kg-day for females (Aida et al., 1992b). Marked suppression of body weight gain was
seen in males and females of the high-dose group. Males and females of the high-dose group
exhibited mild piloerection and emaciation. Relative liver weight was significantly increased in the
mid- and high-dose groups, while relative kidney weight was significantly increased only in the
high-dose groups. At 18 months, dose-dependent reductions
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Table V-5 Summary of Chronic Toxicity Studies for Brominated Trihalomethanes
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromodichloromethane
NTP (1987)
Rat
F344/N
Gavage
(corn oil)
M,
F
50
2 years
(5 d/wk)
0
50 (LOAEL)
100
Renal and hepatic
histopathology
NTP (1987)
Mouse
B6C3F;
Gavage
(corn oil)
M
50
2 years
(5 d/wk)
0
25 (LOAEL)
50
Renal and hepatic
histopathology
NTP (1987)
Mouse
B6C3F;
Gavage
(corn oil)
F
50
2 years
(5 d/wk)
0
75 (LOAEL)
150
Reduced body weight gain
Aida et al.
(1992b); Tobe
et al. (1982)
Rat
Wistar
Diet
M
40
2 years
0
6 (LOAEL)
26
138
Hepatic vacuolization, serum
chemistry
Aida et al.
(1992b); Tobe
et al. (1982)
Rat
Wistar
Diet
F
40
2 years
0
8 (NOAEL)
32 (LOAEL)
168
Hepatic vacuolization, serum
chemistry
Klinefelter et
al. (1995)
Rat
F344
Drinking
water
M
7
1 year
0
22
39 (NOAEL)
No evidence of treatment-related
histopathological or organ
weight effects (see reproductive
effects, section V.E for
additional data from this study)
Dibromochloromethane
Tobe et al.
(1982)
Rat
Wistar
SPF
Diet
M
40
2 years
0
12 (NOAEL)
49 (LOAEL)
196
Serum biochemistry, liver
appearance at necropsy;
decreased body weight gain
Tobe et al.
(1982)
Rat
Wistar
SPF
Diet
F
40
2 years
0
17 (NOAEL)
70 (LOAEL)
278
Serum biochemistry, liver
appearance at necropsy;
decreased body weight gain
NTP (1985)
Rat
F344
Gavage
(corn oil)
M,
F
50
2 years
(5 d/wk)
0
40 (LOAEL)
80
Histologic changes in liver,
including fat accumulation and
ground glass appearance, and
altered basophilic staining
NTP (1985)
Mouse
B6C3F;
Gavage
(corn oil)
M,
F
50
105 weeks
(5 d/wk)
0
50 (LOAEL)
100
Fatty metamorphosis in liver
and follicular cell hyperplasia in
thyroid
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Table V-5 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromoform
Tobe et al.
(1982)
Rat
Wistar
SPF
Diet
M
40
2 years
0
22 (NOAEL)
90 (LOAEL)
364
Enzyme changes and altered
liver appearance at necropsy
Tobe et al.
(1982)
Rat
Wistar
SPF
Diet
F
40
2 years
0
38 (NOAEL)
152 (LOAEL)
619
Enzyme changes and altered
liver appearance at necropsy
NTP (1989a)
Rat
F344/N
Gavage
(corn oil)
M,
F
50
103 weeks
(5 day/wk)
0
100 (LOAEL)
200
Decreased body weight,
lethargy, mild hepatotoxicity
NTP (1989a)
Mouse
B6C3F;
Gavage
(corn oil)
M
50
103 weeks
(5 day/wk)
0
50
100 (NOAEL)
No observed effects on body
weight or hepatotoxicity
NTP (1989a)
Mouse
B6C3F;
Gavage
(corn oil)
F
50
103 weeks
(5 day/wk)
0
100 (LOAEL)
200
Decreased body weight, minimal
to mild fatty changes in liver
* SD, Sprague-Dawley
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in serum cholinesterase activity and increases in y-glutamyl transpeptidase (y-GTP) activity
(indicative of bile duct proliferation) were observed in males, with the changes significant at the
high dose. The mid- and high-dose males also displayed a 27% and 65% reduction, respectively,
in total serum triglycerides (T-Gly) levels when compared to the control group. At 18 months,
serum cholinesterase levels were significantly decreased, while total cholesterol levels were
significantly increased, in all dose groups for the treated females. Serum T-Gly levels (decreased)
and y-GTP activity (increased) were also reported to deviate significantly from control values in
the mid- and high-dose females. The most sensitive markers at 24 months were T-GLY and
serum cholinesterase, with significant changes seen in all of the male treatment groups. Gross
necropsy revealed dose-related yellowing and roughening of the liver surface. Treatment-related
lesions were limited to the liver. At 24 months, fatty degeneration and granuloma were observed
in all dose groups with the exception of granulomas in low-dose females. Specifically, fatty
degeneration and granulomas were observed in low-dose males, but not control males, and fatty
degeneration was observed in low-dose females at a higher rate (8/19) than in control females
(2/32). Cholangiofibrosis was also observed in the high-dose groups. Bile duct proliferation was
observed in most high-dose animals at 6 months, and was prevalent in the controls and all dose
groups by 24 months. Histopathology was observed in all dose groups as early as 6 months with
the exception of low-dose females. Based on the results of Tobe et al. (1982) alone, the NOAEL
was 6 mg/kg-day in males and 8 mg/kg-day in females. The LOAEL was identified as 26 (males)
to 32 (females) mg/kg-day, based on serum enzyme changes and altered liver appearance. Based
on the histopathology data reported for this study by Aida et al. (1992b), however, the entire
study identified a LOAEL of 6 mg/kg-day in male rats and 8 mg/kg-day in female rats.
NTP (1987) administered doses of 0, 50, or 100 mg/kg-day of bromodichloromethane in
corn oil by gavage to male and female F344/N rats (50/sex/dose), 5 days/week for 102 weeks.
The authors observed all animals for clinical signs and recorded body weights (by cage) once per
week for the first 12 weeks of the study and once per month thereafter. A necropsy was
performed on all animals, including those found dead, unless they were excessively autolyzed or
cannibalized. During necropsy, all organs and tissues were examined for grossly visible lesions.
Complete histopathology was performed on all female rats and on high-dose and vehicle-control
male rats. Male rats in the low-dose group that died early in the study were also examined
histologically. Survival of dosed rats was comparable to that of vehicle controls. Mean body
weight of high-dose male and female rats was decreased during the last 1.5 years of the study;
body weight gain of high-dose male and female rats was 86% and 70% of the corresponding
vehicle-control values. Body weight gain of low-dose male and female rats was comparable to
that of the vehicle-control group. No treatment-related clinical signs were observed. In males,
treatment-related nonneoplastic effects included renal cytomegaly, tubular cell hyperplasia,
hepatic necrosis, and fatty metamorphosis. In females, changes included eosinophilic cytoplasmic
change, clear cell change, focal cellular change, fatty metamorphosis of the liver, and tubular cell
hyperplasia of the kidney. Based on these histologic findings, this study identified a LOAEL of
50 mg/kg-day in rats.
NTP (1987) administered bromodichloromethane in corn oil by gavage to male and female
B6C3FX mice (50/sex/dose), 5 days/week for 102 weeks. For males, doses were 0, 25, or
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50 mg/kg-day. For females, doses were 0, 75, or 150 mg/kg-day. Final survival of treated male
mice was comparable to that of vehicle controls. At week 84, survival of female mice was greater
than 50% in all dose groups. After week 84, survival of dosed and vehicle-control female mice
was reduced (final survival: 26/50, 13/50, 15/50 for the 0, 5, and 50 mg/kg-day groups,
respectively), and this decreased survival was associated with ovarian abscesses (8/50, 19/47,
18/49). Body weight gain of high-dose male mice was 87% of that of the vehicle-control group;
the body weight gain of low-dose male mice was comparable to that of the vehicle-control group.
Mean body weight of high-dose female mice was decreased during the last 1.5 years of the study.
The body weight gain was reduced 55% compared to the controls at the high dose and by 25%
among low-dose females. In males, treatment-related nonneoplastic changes included fatty
metamorphosis of the liver, renal cytomegaly, and follicular cell hyperplasia of the thyroid gland.
In females, hyperplasia of the thyroid gland was observed. This study identified a LOAEL of
25 mg/kg-day, based on histopathological findings in male mice.
Klinefelter et al. (1995) reported interim (52-week) necropsy data from a cancer bioassay
in which male F344 rats were administered average concentrations of 0, 330, or 620 mg/L
bromodichloromethane in drinking water. Corresponding doses of 0, 22, and 39 mg/kg-day were
calculated by the authors using water consumption and body weight data. For the interim
sacrifice, 7 animals per dose group were killed, and the testis, epididymis, liver, spleen, kidney,
thyroid, stomach, intestine, and bladder were evaluated histopathologically.
Bromodichloromethane had no effect on body weight or on the kidney, liver, spleen, or thyroid
weight. There was no histopathological evidence of bromodichloromethane-related noncancer or
cancer effects on any of the examined organs. High levels of nephropathy and interstitial cell
hyperplasia were observed, but these lesions were not treatment-related. The NOAEL and
LOAEL for this study are based on reproductive endpoints. These reproductive effects are
summarized in Section V.E. 1.
2. Dibromochloromethane
Tobe et al. (1982) evaluated the chronic effects of dibromochloromethane administered in
the diet to male and female Slc:Wistar SPF rats (40/sex/group) for 24 months. The animals were
5 weeks old at the start of the test and weighed approximately 100 g. Dibromochloromethane
was microencapsulated, and an appropriate amount was mixed with powdered feed. Control
groups (70 rats/sex) received microcapsules without test compound. Body weight and food
consumption were monitored weekly for the first 6 months, every 2 weeks from 6 to 12 months,
and every 4 weeks during the second year of the study. Data were reported from the sacrifices of
9 animals/sex/control group and 5/sex/dose group at 18 months; all surviving animals were
sacrificed at 24 months. Necropsies, hematology, and serum biochemistry were conducted at the
time of sacrifice. No histopathology data for dibromochloromethane have been published from
this study. Dibromochloromethane was administered at dietary levels of 0.0%, 0.022%, 0.088%,
or 0.35%. Based on reported body weights (150 to 475 g for males and 100 to 215 g for
females) and food consumption (15 to 20 g/day for males and 10 to 15 g/day for females), these
levels corresponded to doses of approximately 0, 12, 49, and 196 mg/kg-day for males and 0, 17,
70, and 278 mg/kg-day for females. Marked suppression of body weight gain was seen in males
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and females at the high dose, and mild suppression of body weight gain (about 10%) was seen in
males and females at the mid dose. Decreased T-GLY and serum cholinesterase activity and
increased y-GTP were seen in the mid- and high-dose males and females. Yellowing of the liver
surface was noted in the mid- and high-dose groups, and roughening of the liver surface was
noted in high-dose males. This study suggests NOAELs of 12 mg/kg-day (males) and 17 mg/kg-
day (females), and LOAELs of 49 mg/kg-day (males) and 70 mg/kg-day (females), based on
serum biochemistry data, decreased body weight, and gross necropsy findings.
NTP (1985) investigated the chronic oral toxicity of dibromochloromethane in male and
female F344/N rats. Groups of 50 animals/sex/dose were administered doses of 0, 40, or
80 mg/kg-day by gavage in corn oil for 5 days/week for 104 weeks. Survival was comparable in
all dose groups. Body weight gain was decreased in high-dose males after week 20; final weight
gain was 88% of the control value. Females in both dose groups gained more weight than did the
controls. Histologic lesions in the liver were observed in both males and females at both dose
levels. Changes included fat accumulation, "ground glass" appearance of the cytoplasm, and
altered basophilic staining. This study identified a LOAEL of 40 mg/kg-day for
dibromochloromethane in rats.
NTP (1985) performed a similar chronic oral exposure study of dibromochloromethane
toxicity in male and female B6C3F, mice. Groups of 50 animals/sex/dose were administered
doses of 0, 50, or 100 mg/kg-day by gavage in corn oil for 5 days/week for 105 weeks. Survival
in females was not different from controls, while survival in high-dose males was significantly
decreased. An overdosing accident at week 58 killed 35/50 male mice in the low-dose group, and
this group was not considered further. Mean body weight was decreased in high-dose males and
females, but not in low-dose females. Treatment-related hepatocytomegaly and focal necrosis
were observed in livers of high-dose males. Females showed liver calcification at the high dose
and fatty metamorphosis at both the low and high doses. An increased incidence of follicular cell
hyperplasia in the thyroid was observed in low- and high-dose females relative to the control.
Thyroid lesions were not observed in treated males. This study identified a LOAEL of 50 mg/kg-
day for dibromochloromethane in mice.
3. Bromoform
Tobe et al. (1982) evaluated the chronic effects of bromoform administered in the diet to
male and female Slc:Wistar SPF rats (40/sex/group) for 24 months. The animals were 5 weeks
old at the start of the test and weighed approximately 100 g. Bromoform was microencapsulated,
and administered at dietary levels of 0.0%, 0.04%, 0.16%, or 0.65%. Control groups
(70 rats/sex) received microcapsules without test article. Body weights and food consumption
were monitored weekly for the first 6 months, every 2 weeks from 6 to 12 months, and every
4 weeks during the second year of the study. Data were reported from the sacrifices of
9 animals/sex in the control group and 5/sex/dose in the exposure groups at 18 months; all
surviving animals were sacrificed at 24 months. At each time of sacrifice, necropsies,
hematology, and serum biochemistry were conducted. No histopathology data for bromoform
have been published from this study. Based on reported body weights (150 to 475 g for males
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and 100 to 215 g for females) and food consumption (15 to 20 g/day for males and 10 to 20 g/day
for females), these levels corresponded to doses of about 0, 22, 90, and 364 mg/kg-day for males
and 0, 38, 152, and 619 mg/kg-day for females. Marked suppression of body weight gain was
seen in males and females at the high dose, and mild suppression of body weight gain (about 15%)
was seen in males and females at the mid dose. Dose-related decreases in non-esterified fatty
acids were observed in all treated males and in females at the mid and high dose. Females also
exhibited a dose-related increase in levels of y-GTP with the increases significant at the mid and
high dose. Other serum biochemistry changes in the high-dose groups included decreased serum
triglyceride (T-GLY) and increased AST and ALT activity. Specifically, T-GLY levels
significantly decreased by 86% and 80% in the male and female high-dose groups, respectively, by
study termination. AST and ALT activities at study termination were significantly increased 1.6
to 2.6-fold in animals at the high dose compared to controls, with the exception that the increase
in AST activity in males was statistically nonsignificant. Yellowing and small white spots, and
roughening of the surface were seen in the livers of the mid- and high-dose animals. Roughening
of the liver surface was observed in the high-dose groups. Based on the necropsy findings and the
serum biochemistry data, this study indicated NOAELs of 22 mg/kg-day for males and 38 mg/kg-
day for females, and LOAELs of 90 mg/kg-day for males and 152 mg/kg-day for females.
NTP (1989a) exposed male and female F344/N rats (50/sex/group) to bromoform by
gavage in oil for 103 weeks (5 days/week) at doses of 0, 100, or 200 mg/kg-day. Animals were
observed for clinical signs throughout the study (2 days/week). At termination, necropsy and
histopathological examination were performed on all animals. Body weight gain was decreased
by 37%) in high-dose females and by 29% in high-dose males relative to the respective controls.
Survival of the high-dose males was also decreased. Both males and females were lethargic.
Hepatic fatty change and chronic inflammation were noted in both males and females at both
doses, and minimal necrosis was increased in high-dose males. Nonneoplastic changes were not
reported in other tissues. This study identified a LOAEL of 100 mg/kg-day in both male and
female rats.
NTP (1989a) exposed groups of 50 male B6C3FX mice by gavage in oil to doses of 0, 50,
or 100 mg/kg-day of bromoform for 103 weeks (5 days/week). Groups of 50 female mice were
administered doses of 0, 100, or 200 mg/kg-day. Animals were observed for clinical signs
2 days/week throughout the study. At termination, all animals were necropsied, and a thorough
histological examination of tissues was performed. Decreased survival was observed in females,
but not males. This was at least partly due to a utero-ovarian infection. No clinical signs were
noted. Body weight gains were 82% and 72% of the control values for low- and high-dose
females, respectively, but body weight gain was not affected in males. Increased incidences of
minimal to mild fatty changes were noted in the livers of both low- and high-dose females, but not
males. Nonneoplastic changes were not detected in other tissues. This study identified a LOAEL
of 100 mg/kg-day for female mice, based on decreased body weight and fatty changes of the liver.
No NOAEL for females was identified. For males, a NOAEL of 100 mg/kg-day was identified.
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E. Reproductive and Developmental Effects
Studies that have examined the reproductive and developmental toxicity of the brominated
trihalomethanes are summarized in Table V-9 at the end of this section.
1. Bromodichloromethane
a. Studies in Rats
Ruddick et al. (1983) investigated the teratogenicity and developmental toxicity of
bromodichloromethane in Sprague-Dawley rats. Pregnant dams (15/dose group) were
administered 0, 50, 100, or 200 mg/kg-day by gavage in corn oil on gestation days (GD) 6 to 15.
Body weights were measured on GD 1, on GD 1 through GD 15, and before and after fetuses
were removed by caesarean section on GD 22. On GD 22, females were sacrificed and body
tissues (including the uterus) were removed for pathological examination. Females were
evaluated for the number of resorption sites, and number of fetuses. Maternal blood samples
were collected and evaluated for standard hematology and clinical chemistry parameters. The
liver, heart, brain, spleen, and one kidney were weighed. Standard histopathology was conducted
on control and high dose females (5/group). All fetuses were individually weighed, and evaluated
for viability and external malformations. Histopathologic examination was performed on two
pups per litter. Of the remaining live fetuses, approximately two-thirds were examined for
skeletal alterations and one-third for visceral abnormalities.
Although 15 inseminated females per dose group were exposed to bromodichloro-
methane, not all females became pregnant and/or delivered litters. Therefore, the number of litters
per dose group ranged from 9 to 14. One animal died in the control group, but no deaths
occurred in any of the exposed groups. In the high-dose group, maternal weight gain was
significantly depressed by 38% as compared with controls. Although maternal weight gains were
also reduced in the low- and mid-dose groups (13% and 15%, respectively, as compared with
controls), these differences were not reported as statistically significant. Relative maternal liver
weight was significantly increased in all exposed groups (110%, 110%, and 117% for the low-,
mid-, and high-dose groups, respectively as compared with control values). Relative kidney and
brain weights were also statistically increased in the high-dose group only. These increases in
relative organ weights may have been associated with the decreased body weight gains in treated
females. No treatment-related changes in hematology, clinical chemistry, histopathology, number
of resorptions, and the number of fetuses per litter were noted. No differences between treated
and control groups were reported for fetal weights, gross malformations (terata), and visceral
abnormalities. However, an increase in the incidence of sternebral anomalies was observed in all
treated groups. The number of affected fetuses/number of affected litters were 2/2, 8/4, 9/7, 10/6
for the control, low-, mid-, and high-dose groups, respectively. Statistical significance of
fetotoxic endpoints was not reported by the study authors. An independent statistical analysis
(using the Fisher Exact test) was conducted on the published data for development of this Criteria
Document and demonstrated that none of these increases differed significantly from control values
(p>0.05). A trend test showed a statistically significant dose-related trend (p=0.03); stepwise
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analysis indicated that the trend became nonsignificant if the high-dose (200 mg/kg-day) was
omitted from the analysis. These findings suggest that the LOAEL and NOAEL for
developmental toxicity are 200 and 100 mg/kg-day, respectively. However, it should be noted
that the small sample sizes (the sampling unit is the litter) limited the statistical power of the
experiment to detect possible significant differences at lower doses. Based on significantly
decreased maternal body weight gain, the LOAEL and NOAEL for maternal toxicity are 200 and
100 mg/kg-day, respectively.
Klinefelter et al. (1995) evaluated the effects of bromodichloromethane exposure on male
reproduction during a chronic cancer bioassay study in which F344 rats were administered
bromodichloromethane in drinking water at concentrations of 0, 330, or 620 mg/L. The authors
estimated the doses to be 0, 22, and 39 mg/kg-day. At 52 weeks, the authors conducted an
interim sacrifice, which included an evaluation of epididymal sperm motion parameters and
histopathology of the testes and epididymides. No histologic alterations were observed in any
reproductive tissue. Sperm velocities (mean straight-line, average path, and curvilinear),
however, were significantly decreased at 39 mg/kg-day. No effect on sperm motility was
observed at 22 mg/kg-day. The NOAEL and LOAEL for reproductive effects are thus 22 and 39
mg/kg-day, respectively.
Narotsky et al. (1997) examined both the developmental toxicity and the effect of dosing
vehicle on the developmental toxicity of bromodichloromethane. F344 rats (12 to 14/group) were
administered bromodichloromethane by gavage, in either corn oil or an aqueous vehicle
containing 10% Emulphor®, at dose levels of 0, 25, 50, or 75 mg/kg-day on GD 6 to 15. Dams
were allowed to deliver naturally, and pups were evaluated postnatally. Maternal body weights
were assessed on GD 5, 6, 8, 10, 13, and 20, and all rats were observed for clinical signs of
toxicity throughout the test period. Postnatal day (PND) 1 was defined as GD 22 irrespective of
the actual time of parturition. All pups were examined externally for gross malformations and
weighed on PND 1 and 6. Skeletal and visceral anomalies in the pups were not evaluated.
Following PND 6 examination, the dams were sacrificed and the number of uterine implantation
sites per female was recorded. The uteri of females that did not deliver litters were stained and
evaluated histopathologically to detect any cases of full-litter resorption (FLR). In order to
compare the kinetics of dosing vehicles, a separate experiment was conducted in which pregnant
females (3 to 4 animals per vehicle per time point) were administered a single dose of 75 mg/kg
on GD 6 and whole blood samples were collected at 30 minutes, 90 minutes, 4.5 hours, or 24
hours postdosing. Following blood collection, the animals were sacrificed, blood concentrations
of bromodichloromethane were measured, and pregnancy status was confirmed at necropsy.
In the developmental toxicity study, one animal that received 75 mg/kg-day in corn oil
died before study termination. In the mid- and high-dose groups, clinical signs of toxicity were
evident among animals administered bromodichloromethane in either dosing vehicle. At
75 mg/kg-day, kyphosis (humpback) was observed in animals receiving the oil vehicle, and
piloerection was observed in animals receiving either vehicle. At 50 mg/kg-day, piloerection was
observed in animals receiving the aqueous gavage, and chromodacryorrhea/lacrimation was
observed in animals receiving the oil gavage. Maternal weight gain was significantly decreased in
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all dosed groups receiving the aqueous vehicle and in the 50 and 75 mg/kg-day groups in animals
receiving the oil vehicle on GD 6 to 8 (data not reported for other time periods). Although
maternal weight gain was also reduced at 25 mg/kg-day in animals given the oil vehicle, this
decrease was not statistically significant. However, a two-way analysis of variance (ANOVA)
indicated that there was no interaction between vehicle and dose for this maternal endpoint. All
control and 25 mg/kg-day litters survived the test period; however, FLR was observed at 50 and
75 mg/kg-day with both dosing vehicles. Statistical analysis (ANOVA) of FLR incidence showed
a significant vehicle-dose interaction. For females receiving bromodichloromethane in corn oil,
FLR was reported in 8 and 83% of the litters at 50 and 75 mg/kg-day, respectively; an additional
high-dose litter was carried to term but was delivered late (GD 23), and all pups died by PND 6.
For females receiving the aqueous vehicle, FLR was observed in 17 and 21% of the litters at 50
and 75 mg/kg-day, respectively. There were no effects on gestation length, pre- or postnatal
survival, or pup morphology in surviving litters, with the exception noted above in the 75 mg/kg-
day oil vehicle group. Based on full litter resorption, the LOAEL for developmental toxicity is 50
mg/kg-day for both vehicles, and the corresponding NOAEL is 25 mg/kg-day. Based on
significantly reduced body weight gain during GD 6 to 8 in dams receiving the aqueous vehicle,
the LOAEL for maternal toxicity is the lowest dose tested, 25 mg/kg-day, and a NOAEL could
not be determined.
Analysis of bromodichloromethane concentrations in blood indicated that circulating levels
decreased over time with both vehicles, but tended to be higher following corn oil administration.
Bromodichloromethane blood concentrations were thus vehicle-dependent and differed
statistically at both 4.5 and 24 hours postdosing (mean of 3.1 ng/mL versus 0.4 ng/mL for oil and
aqueous vehicles, respectively, at 24 hours). The elimination half-life of bromodichloromethane
was estimated to be 3.6 hours when administered in corn oil and 2.7 hours when given in the
aqueous vehicle.
Narotsky et al. (1997) also calculated both an ED05 (i.e., the effective dose producing a
5% increase in response rate above background) and a benchmark dose (BMD; as defined by the
authors, the BMD is the lower confidence interval of the ED05) for each vehicle. For the corn oil
vehicle, the ED05 and BMD were 48.4 and 39.3 mg/kg-day, respectively. For the aqueous
vehicle, the ED05 and BMD were 33.3 and 11.3 mg/kg-day, respectively. The study authors noted
that the greater BMD value for the corn oil vehicle seemed counterintuitive in view of the higher
FLR response rate in the 75 mg/kg-day aqueous vehicle group (83% for aqueous vehicle versus
21% for corn oil vehicle). However, the dose response for bromodichloromethane-induced FLR
differed markedly between vehicles, and the response rate in the 50 mg/kg-day corn oil vehicle
group (8%>) closely approximated 5%, the effect level defined by the ED05. According to the
study authors, this resulted in a smaller confidence interval around the ED05 for the corn oil
vehicle, yielding a less conservative (i.e., higher) BMD. These findings are consistent with the
pharmacokinetic data demonstrating a slower elimination of bromodichloromethane following a
single dose of 75 mg/kg in corn oil as compared with the same dose in aqueous vehicle, and
suggest that the influence of vehicle on FLR rate is dose-dependent.
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NTP (1998) conducted a short-term reproductive and developmental toxicity screen in
Sprague-Dawley rats to evaluate the effects of bromodichloromethane administered in drinking
water. The study was designed to identify the physiological endpoints most sensitive to
bromodichloromethane exposure, and assessed development, female reproduction, male
reproduction, hematology, clinical chemistry, and pathology. In males, the reproductive
endpoints evaluated included testis and epididymis weight, sperm morphology, density and
motility. The female reproductive parameters evaluated included mating index, pregnancy index,
fertility index, gestation index, number of live births, number of resorptions, implants per litter,
corpora lutea and pre-and post- implantation loss. Concentrations of 0, 100, 700 and 1300 ppm
bromodichloromethane were selected for use in this study based on decreased water consumption
observed in a preliminary 14-day range-finding study (see Section V.B. 1). Two groups of male
Sprague-Dawley rats and three groups of female Sprague-Dawley rats were assigned to treatment
groups as indicated in Table V-6.
Table V-6 NTP (1998) Study Design
Gender
Group
Description
# Animals per Dose Group
0 ppm*
100 ppm
700 ppm
1300 ppm
Male
A
non-BrdU treated
10
10
10
10
B
BrdU treated
5
5
5
8
Female
A
peri-conception exposure
10
10
10
10
B
gestational exposure
13
13
13
13
C
BrdU treated, peri-
conception exposure
5
5
5
8
* Control animals received deionized water
Test animals were dosed for 25 to 30 days, with the exception of Group B females which
were dosed from GD 6 to evidence of littering/birth (total duration approximately 15 to 16 days).
Based on measured water consumption, the nominal concentrations of 0, 100, 700 and 1300 ppm
were equivalent to doses of 0, 8, 41, and 68 mg bromodichloromethane/kg-day for all male rats
and 0, 14, 72 and 116 mg bromodichloromethane/kg-day for all female rats in groups A and C.
The calculated doses for Group B females were 0, 13, 54, and 90 mg/kg-day. All animals
survived the treatment period, with the exception of one Group A male in the 700 ppm dose
group. Body weight and food and water consumption were decreased at many time points for
animals dosed with 700 and 1300 ppm bromodichloromethane. Body weights in the dosed
groups were decreased from 5% to 13%, food consumption was decreased from 14% to 53%,
and water consumption was decreased from 7% to 86% relative to control animals. However,
bromodichloromethane exposure did not alter any reproductive parameter investigated in males or
females, with the exception of a non-dose-related increase in the number of live fetuses per birth
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at the 100 ppm concentration in Group C females, and a slight decrease in the number of corpora
lutea at the 700 ppm concentration in Group A females. On the basis of these results, NTP
(1998) concluded that bromodichloromethane was not a short-term developmental or
reproductive toxicant any of the doses tested in the study. The reproductive/developmental
NOAELs are 68 and 116 mg/kg-day for male and female rats, respectively. The adult NOAEL
and LOAEL for this study were identified on the basis of hepatic effects, which are discussed in
detail in Section V.B. 1.
Bielmeier et al. (2001) conducted a series of experiments to investigate the mode of action
in bromodichloromethane-induced full litter resorption (FLR). These experiments are
summarized in Table V-7. This series of experiments included a strain comparison of F344 and
Sprague-Dawley (SD) rats, a critical period study, and two hormone profile studies. In the strain
comparison experiment, female SD rats (13 to 14/dose group) were dosed with 0, 75, or 100
mg/kg-day by aqueous gavage in 10% Emulphor® on GD 6 to 10. F344 rats (12 to 14/dose
group) were concurrently dosed with 0 or 75 mg/kg-day administered in the same vehicle. The
incidence of FLR in the bromodichloromethane-treated F344 rats was 62%, while the incidence of
FLR in SD rats treated with 75 or 100 mg/kg-day of bromodichloromethane was 0%. Both
strains of rats showed similar signs of maternal toxicity, and the percent body weight loss after the
first day of dosing was comparable for SD rats (no resorption observed) and the F344 rats that
resorbed their litters. F344 rats that maintained their pregnancies generally did not lose weight
after the first dose, although they did experience significantly less weight gain than the controls.
Both strains of rats had similar incidences of piloerection. However, the strains showed different
ocular responses to compound administration. One half (7/14) of the treated F344 rats showed
lacrimation and/or excessive blinking shortly after dosing during the first two days of compound
administration. In comparison, only 1/28 of the SD rats exhibited this response. The study
authors reported that lacrimation was not predictive of FLR in F344 rats. The rats were allowed
to deliver and pups were examined on postnatal days 1 and 6. Surviving litters appeared normal
and no effect on post-natal survival, litter size, or pup weight was observed.
Bielmeier et al. (2001) conducted a second experiment to identify the critical period for
bromodichloromethane-induced FLR. Two different five day periods during organogenesis were
compared. F344 rats (12 to 13/dose group) were dosed with 75 mg/kg-day by gavage in 10%
Emulphor® on GD 6 to 10 (which includes the luteinizing hormone-dependent period of
pregnancy) or GD 11 to 15 (a luteinizing hormone-independent period). Rats (8 to 10/dose
group) dosed with 0 or 75 mg/kg-day on GD 6 to 15 served as negative and positive controls,
respectively. FLR occurred only in rats treated on GD 6 to 10 or GD 6 to 15 (incidences of 75%
and 50%), respectively). In contrast, all rats treated with bromodichloromethane on GD 11 to 15
maintained their litters. Surviving litters appeared normal and no effect on post-natal survival,
litter size, or pup weight was observed. This finding was interpreted by the study authors as
evidence for an effect of bromodichloromethane on luteinizing hormone secretion or signal
transduction.
To investigate possible endocrine responses to bromodichloromethane treatment that
might be associated with FLR, Bielmeier et al. (2001) characterized the serum profiles of
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luteinizing hormone (LH) and progesterone in two experiments. Progesterone is necessary for the
maintenance of pregnancy, while LH participates in the maintenance of the corpora lutea which
secrete progesterone. In the first experiment, F344 rats (8-10/dose) were given a 100
Table V-7 Summary of Experiments Conducted by Bielmeier et al. (2001)
Study/Strain
Dose
(mg/kg-day)
Treatment
Period
Number of animals
% FLR
Treated
Pregnant
Resorbed
Strain Comparison
F344
0
GD 6-10
12
11
0
0
F344
75
GD 6-10
14
13
8
62**
SD
0
GD 6-10
13
13
0
0
SD
75
GD 6-10
14
14
0
0
SD
100
GD 6-10
14
14
0
0
Critical Study Period
F344
0
GD 6-15
8
8
0
0
F344
75
GD 6-15
10
10
5
50*
F344
75
GD 6-10
12
12
9
75**
F344
75
GD 11-15
13
13
0
0
Hormone Profile P
F344
0
GD 8-9
8
7
0
0
F344
100
GD 8
10
10
6
60*
F344
100
GD 9
10
9
9
100***
Hormone Profile IIb
F344
0
GD 9
8
8
0
0
F344
75
GD 9
11
11
7
64*
F344
100
GD 9
10
10
9
90***
Source: Table 1 in Bielmeier et al. (2001)
Abbreviations: GD, gestation day; FLR, Ml litter resorption; SD, Sprague-Dawley
" Tail blood collected once daily on GD 9 to 12.
b Tail blood collected at 0, 6, 12, and 24 hours after dosing.
* p<0.05; ** p<0.01; *** p<0.001for significant differences from controls (Fisher's Exact Test).
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mg/kg dose by gavage in 10% Emulphor® on GD 8 or 9. Tail blood samples were collected once
daily on GD 9 through 12 and progesterone and LH were determined by ELISA and
radioimmunoassay, respectively. As reported by the study authors, the first blood collection (GD
9) established baseline levels for the GD 9-treated group and 24-hour post-dosing levels for
animals treated on GD 8. FLR occurred in 0, 60, and 100% of the control, GD 8 and GD 9 dose
groups, respectively. When progesterone was measured 24 hours after dosing, all rats that
resorbed their litters had markedly reduced levels of progesterone when compared to the control
rats or to rats that retained their litters. The mean progesterone (± standard error) levels in
animals dosed on GD 9 (n = 9) that resorbed their litters decreased from a baseline level of 137.94
± 11.44 ng/mL to 48.45 ± 23.57 ng/mL within 24 hours. The mean progesterone levels 24 hours
post-treatment for animals treated on GD 8 was 67.01 ± 16.22 ng/mL for the animals that
resorbed their litters (n = 6), while the corresponding control group mean was 127.19 ± 14.89
ng/mL (n=7). Three days after dosing the progesterone levels in the resorbed groups were
reported to be comparable to those of two nonpregnant animals in the study. In contrast to the
changes in progesterone levels, LH levels were unaffected in treated rats when measured 24 hours
after dosing. Serum LH levels were elevated on GD 11 to 12 in the resorbed groups. In the
groups of animals that resorbed their litters, the mean serum levels of LH rose from about 0.20
ng/mL on GD 10 to about 0.80 ng/mL on GD 11 and remained elevated on GD 12. These levels
were reported to be comparable to the levels observed in two nonpregnant animals in the study.
During the same time period, the control levels dropped from 0.31 ng/mL to 0.14 ng/mL. Litters
delivered to dams exposed to bromodichloromethane had significantly heavier pups and fewer
pups than control litters.
In the second hormone profile study conducted by Bielmeier et al. (2001), doses of 0, 75,
or 100 mg/kg-day were administered to rats (8-11/dose) on GD 9 by gavage in 10% Emulphor®.
Blood samples were collected at 0, 6, 12, and 24 hours after dosing. The dose-related incidence
of FLR was 0% (0/8), 64% (7/11), and 90% (9/10), respectively. The progesterone levels in all
groups peaked at six hours. The peak level observed in dams administered 75 mg/kg that
resorbed their litters was significantly reduced when compared to the controls. The peak in
animals receiving 100 mg/kg was slightly reduced, but the effect did not reach statistical
significance. At 12 hours, the progesterone levels of dams that resorbed their litters were
significantly reduced in both bromodichloromethane dose groups. At 24 hours, progesterone
levels were further reduced in dams that displayed FLR. The mean progesterone levels in dams
that retained their litters were comparable to the controls. LH levels were comparable in the
control and treatment groups at all tested time points. When analyzed by a repeated measures
procedure, the results indicate a significant decline among all groups over the 24 hour period.
However, no significant differences were noted between groups. In contrast to the results for pup
number in the previous hormone profile experiment, litters of treated dams in this experiment had
significantly more pups than controls.
The series of experiments conducted by Bielmeier et al. (2001) identified a LOAEL of 75
mg/kg-day (the lowest dose tested) based on FLR in F344 rats. A NOAEL was not identified.
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The Chlorine Chemistry Council sponsored a range finding reproductive toxicity study of
bromodichloromethane in rats (CCC, 2000c), which was conducted according to standard U.S.
EPA test guidelines (U.S. EPA, 1998c) and GLP standards. This study is summarized in
Christian (2001b). Male and female Sprague Dawley rats (10/sex/group) were randomly assigned
to five exposure groups. Additional rats (6 males/group and 15 females/group) were assigned to
satellite groups for collection of samples for analysis of bromodichloromethane concentrations in
selected tissues and fluids (see Section III.B). Bromodichloromethane was administered to
parental rats (P generation) in drinking water at concentrations of 0, 50, 150, 450, or 1350 ppm.
Exposure began 14 days before cohabitation and continued until the day of sacrifice. Female
estrous cycle evaluations were performed daily, beginning 14 days before exposure initiation and
continuing for 14 days after the first day of exposure. Clinical observations were recorded daily
during the exposure period. Male body weights were recorded weekly during the entire exposure
period and at sacrifice; female body weights were recorded weekly during precohabitation and
cohabitation, on GD 0, 7, 14, 21, and 25, and on lactation days (LD) 1, 5, 8, 11, 15, 22, and 29.
Lactation was extended for one week (LD 22-29) beyond the normal 3-week period because F,
pup body weights in the three highest dose groups were significantly reduced on LD 21 relative to
control values (results are described below). Water and feed consumption were recorded weekly
and at sacrifice for males during the entire exposure period (except for feed consumption during
cohabitation), and more frequently for females during gestation and lactation. On LD 29, two F,
pups per sex were selected from each litter for an additional week of postweaning observation,
provided ad libitum access to water containing the same concentration of bromodichloromethane
administered to their parents (P generation), and sacrificed on Day 8 postweaning. P generation
female rats were assessed for duration of gestation, fertility index, gestation index, number and
sex of offspring per litter, number of implantation sites, and clinical signs of toxicity during the
postpartum period. During lactation, maternal behavior was observed and recorded on LD 1, 5,
8, 11, 22, and 29. Litters were externally examined following delivery to identify the number and
sex of pups, stillbirths and live births, and gross external malformations. Litters were observed at
least twice daily during the preweaning and postweaning period for pup deaths and clinical signs
of toxicity. Litter size and viability, viability indices, lactation indices, percent survival, and sex
ratios were calculated. During the postweaning period of observation, body weights and feed
consumption were recorded at weaning and on day 8 postweaning; water consumption was
recorded daily.
At the end of the parental exposure periods (64 days for males and a maximum of 74 days
for females), all P generation rats were sacrificed and a gross necropsy of the thoracic, abdominal,
and pelvic viscera was performed. In addition, testes and epididymides were excised from males
and paired organ weights were measured. F, pups exposed to bromodichloromethane in their
drinking water for one week following weaning were sacrificed on Day 8 postweaning and
examined for gross lesions. No histopathology was performed on either the P or F, generation.
The consumption of bromodichloromethane was calculated from measured water intake
and measured concentrations of the test article. Mean consumed dosages of bromodichloro-
methane for P generation male rats during the entire exposure period, P generation female rats
during different physiologic stages, and F, postweaning rats are summarized in Table V-8. Males
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and nonpregnant female rats tended to consume similar amounts of bromodichloromethane.
Progressively higher dosages were consumed by female rats in the pre-mating, gestation, and
lactation periods, respectively. The highest dosages among all groups were consumed by Fx
female rats during the one-week postweaning observation period. A possible source of error in
the estimates for lactating females was consumption of the dams' drinking water by their pups.
In the P generation, all male rats and all females except one survived to scheduled
sacrifice. Exposure-dependent reductions in both absolute (g/day) and relative (g/kg body
weight-day) water consumption were observed in all rats of both sexes and were attributed to
Table V-8 Mean Consumed Doses (mg/kg-day) of Bromodichloromethane in the Range
Finding Study Conducted by CCC (2000c)
Gen.
Sex
Exposure Interval
0 ppm
50 ppm
150 ppm
450 ppm
1350 ppm
P
M
Full study
Study Days 1-64
0.0
4.2 ±0.4
11.8 ± 1.8
27.5 ±3.4
67.2 ±5.6
P
F
Pre-mating
Study days 1-15
0.0
4.7 ±0.8
13.3 ±2.0
23.5 ±5.3
70.8 ± 1.8
P
F
Gestation days 0-21
0.0
5.4 ±0.7
16.3 ±2.2
41.7 ±6.4
111.7 ± 6.2
P
F
Lactation days 1-15
0.0
11.0 ± 1.9
31.4 ±2.6
90.3 ±7.3
222.4 ± 19.9
F,
M
Postweaning days 1-8
0.0
13.6 ±3.5
41.4 ±7.1
106.9 ±20.8
297.8 ± 113.8
F,
F
Postweaning days 1-8
0.0
13.9 ±2.6
40.1 ±6.8
117.9 ±42.7
333.6 ± 110.6
taste aversion. Reduced water consumption was most pronounced during the first week of
exposure, and was evident during precohabitation and cohabitation in both sexes, and during post-
cohabitation in males and gestation in females. However, the decrease in water consumption
during these times was not as severe as that observed during the first week of exposure.
Decreased water consumption was not clearly noted in females during lactation, presumably
reflecting the physiologic demands for high fluid consumption during this period. Exposure-
related decreases in feed consumption were noted for males and females in the 150, 450, and
1350 ppm exposure groups, and persisted in the 450 and 1350 ppm females during gestation and
lactation. Treatment-related clinical signs of toxicity were observed in both sexes in the 1350
ppm exposure groups and were considered to be generally associated with reduced water
consumption. Males exhibited dehydration, emaciation, chromorhinorrhea, and
chromodacryorrhea during the pre-mating, cohabitation and post-cohabitation periods; however,
the most severe symptoms resolved within the first 17 days of exposure. Among females, urine-
stained fur was observed in one or more animals in the three highest dose groups during lactation
and was considered to be treatment-related. Reductions in mean body weight gain and body
weight were observed in male rats in the 450 and 1350 ppm exposure groups relative to controls.
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These effects were most severe during the first week of exposure. Mean body weight gains for
the 450 ppm and 1350 ppm male groups over the entire exposure period were 91.3% and 76.3%
of the control values, respectively. At study termination, mean male body weights were 96.5%
and 91.6 % for the 450 ppm and 1350 ppm, respectively, relative to control values. In female
rats, reductions in body weight gain and body weight occurred in 150, 450, and 1350 ppm
groups. These effects were most severe during the first week of exposure, but also persisted
throughout gestation and lactation. During gestation, the mean reductions in female body weight
in the 150, 450, and 1350 ppm groups were 95.8%, 95.3%, and 85.3% of the control values,
respectively. Mean body weights for the entire lactation period were not presented in the study
report; however, inspection of the data, presented separately for LD 1, 8, 15, 22, and 29,
indicated that female body weights were decreased relative to controls in a dose-dependent
manner in the three highest dose groups at all time points.
No gross lesions attributable to bromodichloromethane were observed in the P generation
male or female rats at necropsy. The absolute paired epididymal weights were slightly reduced
(93.2% and 92.5%, respectively) in the 450 and 1350 ppm exposure groups. However, relative
paired epididymal weights were unaffected, suggesting that the decreased absolute values were
associated with the reduced terminal body weights in these groups. Absolute and relative testes
weights were not altered by exposure to bromodichloromethane. No effects of bromodichloro-
methane were observed on any of the measured reproductive parameters in P generation male or
female rats. However, bromodichloromethane exposure was associated with a concentration-
dependent reduction in F, pup body weights in the 150, 450, and 1350 ppm exposure groups.
Pup weights were reported for postpartum days 1, 5, 8, 15, 22, and 29. The mean litter pup
weights in treated groups were comparable to the mean litter pup weight of the control group on
LD 1. Beginning on LD 5, reductions in mean pup weights in the three highest dose groups
increased with increasing dose and duration of the postpartum period. On LD 29, pup weights
averaged 7, 12, and 29% less than controls in the 150, 450, and 1350 ppm exposure groups,
respectively. Reduced body weight gain continued to occur in the Fx pups administered parental
concentrations of bromodichloromethane in drinking water for one week postweaning. No
reductions in either body weight gain or body weight were observed in F, pup litters in the 50
ppm group during lactation or the one-week postweaning period.
Statistical analysis was not conducted in this range finding study. Based on decreased pup
weight and pup weight gain, the LOAEL for developmental toxicity is 150 ppm, and the
corresponding NOAEL is 50 ppm. Although the effect of reduced water consumption on the
decreases in feed consumption, body weight gain, and body weight observed in the P generation
adults is unclear, the LOAEL for parental toxicity is considered to be 150 ppm and the NOAEL is
50 ppm. Due to the marked changes in drinking water consumption by P generation female rats
during different physiological stages (pre-mating, mating, gestation, and lactation), it is not
possible to convert the administered drinking water concentrations into biologically meaningful
average daily doses.
The Chlorine Chemistry Council sponsored a developmental toxicity study of
bromodichloromethane in rats (CCC, 2000d). Data from this study are summarized in Christian
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et al. (2001a). This study was conducted in accordance with U.S. EPA Health Effects Test
Guidelines OPPTS 870.3700: Prenatal Developmental Toxicity Study (U.S. EPA, 1998c) and
U.S. EPA Good Laboratory Practice Standards (40 CFR Part 160/792). Female Sprague-Dawley
rats (25/exposure group) were exposed to bromodichloromethane in the drinking water at
concentrations of 0, 50, 150, 450, and 900 ppm on days 6 to 21 of gestation (GD 6 to 21). The
rats were examined daily during the exposure period for clinical signs related to exposure,
abortions, premature deliveries and deaths. Body weights, water consumption, and feed
consumption were recorded at intervals throughout the exposure period. All study animals were
sacrificed on GD 21 and caesarean-sectioned. A gross necropsy of the thoracic, abdominal, and
pelvic viscera was performed. Data was collected for gravid uterus weight (with cervix), number
of corpora lutea/per ovary, evidence of pregnancy, number and distribution of implantation sites,
live and dead fetuses, early and late resorption, and placental abnormalities (size, color, or shape).
Individual fetuses were weighed, sexed, and examined for gross external abnormalities.
Approximately one-half of the fetuses in each litter were examined for soft tissue alterations and
the heads of these fetuses were examined by free-hand sectioning. The remaining fetuses in each
litter were examined for skeletal alterations.
Consumed dosages for GD 6 to 21 were calculated from measured water consumption
and measured body weights and averaged 0, 2.2, 18.4, 45.0, and 82.0 mg/kg-day, respectively.
No abortions, premature deliveries, deaths or treatment-related clinical signs were observed
during the study and all rats survived until scheduled sacrifice. No treatment-related gross lesions
were identified at autopsy. Exposure-related decreases in maternal body weight gains occurred in
all groups administered bromodichloromethane in the drinking water on the first day of exposure
(GD 6 to 7). The reduction in maternal body weight gain reached statistical significance in the
150, 450, and 900 ppm groups. The effect was most severe on these days and appeared to be
related to taste aversion. The effect on maternal body weight gain was persistent in the 450 and
900 ppm exposure groups. In contrast, the effect was transient in the 50 and 150 ppm exposure
groups. Average body weights were significantly reduced in the 450 and 900 ppm exposure
groups on GD 7 to 21. Average maternal body weights in the same groups were significantly
reduced at terminal sacrifice when corrected for gravid uterine weight.
Statistically significant, exposure-related decreases in absolute (g/day) and relative (g/kg-
day) water consumption were observed in all groups exposed to bromodichloromethane. This
effect was evident for the entire exposure period (GD 6 to 21) and the entire gestation period
(GD 0 to 21). Within the exposure period, the effects were most pronounced on the first two
days of exposure and gradually decreased in severity with continued exposure. Exposure-related
decreases in absolute and relative feed consumption were observed in the 150, 450, and 900 ppm
groups. In the 150 ppm group, the effects were statistically significant only on GD 12 to 15 and
thus were considered to be of little biological importance by the study authors. In the 450 ppm
and 900 ppm groups, absolute and relative feed consumption was significantly reduced for the
entire exposure period (GD 6 to 21), the entire gestation period (GD 0 to 21), and at many
intervals within the exposure period. The effect of bromodichloromethane on feed consumption
tended to be most severe during the first two days of compound administration.
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Caesarean section and litter parameters were unaffected by exposure of the dams to
bromodichloromethane feed concentrations up to 900 ppm. Litter averages for corpora lutea,
implantations, litter sizes, proportion of live fetuses, early or late resorptions, fetal body weights,
percent reabsorbed conceptuses, and percent live fetuses were comparable among all study groups
and no significant differences were observed. No cases of full litter resorption were observed and
there were no dead fetuses. Late resorption occurred in one control group litter. All placentae
appeared normal. All values for the examined litter parameters were within the historical range of
the test facility (Argus Research Laboratories, Horsham, PA) or litter incidences of any gross
external or soft tissue alterations. With respect to skeletal alterations, no skeletal malformations
were observed in any fetus. The only statistically significant (p<0.01) changes in the occurrence
of skeletal variations were reversible delays in ossification. These included an increased fetal
incidence (fetal incidence: 0 ppm, 1/182; 50 ppm, 0/199; 150 ppm, 0/200; 450 ppm, 0/188; 900
ppm, 4/194; litter incidence: 0 ppm, 1/23; 50 ppm, 0/25; 150 ppm, 0/25; 450 ppm, 0/25; 900 ppm,
2/25) of wavy ribs in the 900 ppm exposure group and a decreased number of ossification sites
per fetus per litter for the forelimb phalanges (Mean number ± SD of ossification sites: 8.14 ±
0.91, 8.30 ± 0.65, 8.09 ± 0.63, 7.92 ± 0.78, 7.46 ± 0.78) and the hindlimb metatarsals (Mean
number ± SD of ossification sites: 4.81 ± 0.25, 4.86 ± 0.23, 4.78 ± 0.27, 4.71 ± 0.28, 4.53 ± 0.33)
and phalanges (Mean number ± SD of ossification sites: 6.20 ± 1.19, 6.20 ± 1.17, 5.84 ± 0.94,
5.86 ± 0.79, 5.29 ± 0.54). The increased fetal incidence of wavy ribs was considered unrelated to
bromodichloromethane exposure by the study authors because the litter incidence (the more
relevant measure of effect) did not differ significantly from the control and was within the
historical range for this alteration at the test facility.
The concentration-based maternal NOAEL and LOAEL for this study were 150 ppm and
450 ppm, respectively, based on statistically significant, persistent reductions in maternal body
weight and body weight gains. Based on the mean consumed dosage of bromodichloromethane,
these concentrations correspond to doses of 18.4 mg/kg-day and 45.0 mg/kg-day, respectively.
The concentration-based developmental NOAEL and LOAEL were 450 ppm and 900 ppm,
respectively, based on a significantly decreased number of ossification sites per fetus for the
forelimb phalanges and the hindlimb metatarsals and phalanges. These concentrations correspond
to mean consumed doses of 45.0 mg/kg-day and 82.0 mg/kg-day, respectively.
b. Studies in Rabbits
The Chlorine Chemistry council sponsored a range-finding developmental toxicity study in
New Zealand White rabbits (CCC, 2000a). The data from this study have been summarized in
Christian et al. (2001b). This study was conducted in accordance with U.S. EPA Health Effects
Test Guidelines OPPTS 870.3700: Prenatal Developmental Toxicity Study (U.S. EPA, 1998c)
and U.S. EPA Good Laboratory Practice Standards (40 CFR Part 160/792).
Bromodichloromethane was provided to New Zealand White presumed pregnant rabbits
(5/group) in the drinking water at concentrations of 0, 50, 150, 450, and 1350 ppm on GD 6 to
29. Additional rabbits (4/group) were similarly assigned to satellite treatment groups for use in
the collection of samples for analysis of tissue concentrations of bromodichloromethane
(discussed in Section III.B). Body weights were recorded on GDs 0, 4, daily during the exposure
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period, and on the day of sacrifice. Feed and water consumption data were recorded daily. The
rabbits were sacrificed on GD 29 and gross necropsy of the thoracic, pelvic, and abdominal
viscera were performed. The gravid uterus was excised and weighed. Examinations were made
for number and distribution of corpora lutea, implantation sites, early and late resorptions live and
dead fetuses. Each fetus was examined for gross external alterations and sex (by internal
examination).
The mean consumed daily doses of bromodichloromethane for GDs 6 to 29 were 0.0, 4.9,
13.9, 32.3, and 76.3 mg/kg-day, as determined from measured body weights and measured water
consumption. Absolute (g/day) and relative (g/kg-day) maternal water intake for the exposure
period was decreased in each group administered bromodichloromethane. The relative
consumption values were 92%, 87%, 67%, and 53% of the control group value, respectively.
Absolute and relative feed consumption values were reduced in a time (onset of reductions
delayed in the 50 and 150 ppm exposure groups) and exposure-dependent manner. The relative
values for feed consumption were 96%, 96%, 90%, and 82% of the control group value for the
exposure period. No deaths, abortions, or premature deliveries occurred during the study. No
treatment-related clinical signs or gross lesions were observed. Maternal body weight gains for
the exposure period were 82%, 80%, 73%, and 50%, respectively, relative to the controls. The
study authors questioned whether these reductions were associated with bromodichloromethane
exposure since similar changes did not occur in the satellite exposure group, and suggested that
the reduced body weight gains were artifacts of the small sample size used in the study. When
body weights were corrected for gravid uterus weight, all exposed groups in the main study
experienced body weight loss while body weight gain occurred in the control group. Absolute
uterine weights were reduced in the 450 and 1350 ppm groups. This finding was most likely
associated with reduced body weight in these groups, since relative gravid uterine weights in all
dosed groups were similar to that of the control.
Litter averages for corpora lutea, implantations, litter sizes, live and dead fetuses, early
and late resorptions, percent dead or resorbed conceptuses, fetal body weights, and percent live
male fetuses were comparable for the control and all exposure groups and within the historical
ranges for the test facility (Argus Laboratories, Horsham, PA). All placentas were normal in
appearance. No gross external fetal alterations were observed in the control or treatment groups.
In the satellite study (described in Section III.B), analytical analyses detected trace amounts of
bromodichloromethane in placental samples from two litters in the 1350 ppm group and in one
fetus from the 1350 ppm group. Bromodichloromethane was not detected in amniotic fluid or
maternal plasma. One litter in the 450 ppm satellite exposure group consisted of only early
resorptions. The concentration-based LOAEL for maternal toxicity in this study is 50 ppm, the
lowest concentration tested, based on reduced body weight gain. This concentration corresponds
to a mean daily dose of approximately 4.9 mg/kg-day. The concentration-based NOAEL for
developmental effects was 1350 ppm (the highest dose tested). This corresponds to a mean daily
intake of approximately 76.3 mg/kg-day.
The Chlorine Chemistry Council (CCC, 2000b) sponsored a developmental toxicity study
in New Zealand White rabbits. Data from this study were summarized in Christian et al. (2001a).
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Bromodichloromethane was provided to pregnant rabbits (25/dose group) at concentrations of 0,
15, 150, 450, and 900 ppm in the drinking water on GD 6-29. Consumed doses were calculated
from measured water intake and measured body weights and averaged 0, 1.4, 13.4, 35.6, and 55.3
mg/kg-day, respectively, over the 14 day treatment period. Feed consumption, water intake, and
body weight were monitored daily during the exposure period. The rabbits were sacrificed on GD
29 and examined for gross lesions of the thoracic, abdominal, and pelvic viscera. Uterine weight,
number of implantation sites, uterine contents, and number of corpora lutea were recorded. Each
fetus was examined for weight, gross external alterations, skeletal alterations, and sex. Visceral
alterations and cavitated organs were evaluated by dissection. One rabbit in the 900 ppm dose
group was sacrificed moribund with hindlimb paralysis caused by a back injury. Another rabbit in
the 900 ppm exposure group had a dead litter as a result of a non-treatment related uterine
abnormality. No treatment-related clinical signs or necropsy results were observed. The 450 and
900 ppm exposure groups had significantly reduced feed and water consumption rates throughout
the exposure period. These groups also had significantly reduced body weight gains and
corrected (for weight of gravid uterus) body weight gains for both the bromodichloromethane
exposure period (GD 6 to 29) and the entire gestation period (GD 0 to 29).
Bromodichloromethane had no observable effect on implantations, corpora lutea, live litter size,
early or late resorptions, percentage of male fetuses, percentage of resorbed conceptuses, or fetal
body weight. The number of litters with any alteration, the number of fetuses with any alteration,
the average percentage of fetuses with any alteration did not differ significantly from the control.
Although statistically significant increases in the number of fused sterna centra were observed in
the 150 and 450 ppm groups, this effect was not dose-related and the observed incidences were
within the historical range for the testing facility. Litter averages for ossification sites per fetus
did not differ significantly from the control and were within historical range for the testing facility.
The NOAEL and LOAEL identified for maternal toxicity in this study were 13.4 mg/kg-day (150
ppm) and 35.6 mg/kg-day (450 ppm), respectively, based on decreased body weight gain. The
developmental NOAEL was 55.3 mg/kg-day (900 ppm) based on absence of statistically
significant, dose-related effects at any tested concentration.
Christian et al. (2002) summarized the results of a two-generation reproductive toxicity
study on bromodichloromethane conducted in Sprague-Dawley rats. The study was sponsored by
the Chlorine Chemistry Council (CCC, 2002) and was conducted in accordance with U.S. EPA
Health Effects Test Guideline OPPTS 870.3800: Reproduction and Fertility Effects (U.S. EPA,
1998b) and U.S. EPA Good Laboratory Practice Standards (40 CFR Part 160/792).
Bromodichloromethane was continuously provided to test animals in the drinking water at
concentrations of 0, 50, 150, or 450 ppm. Drinking water solutions were prepared at least once
weekly and precautions were taken to prevent contamination of the solutions by extraneous
sources of chlorine. Concentrations were verified analytically at the beginning and end of each
exposure period. The tested concentrations were selected on the basis of results obtained in the
developmental toxicity screening study conducted by NTP (1998) and data obtained in a
range-finding study (CCC, 2001c; Christian et al., 2001b). Exposure of the parental generation
(30 rats/sex/concentration) was initiated when the test animals were approximately 43 days of age
and continued through a 70-day pre-mating period and a cohabitation period of up to 14 days.
Parental generation males were exposed for approximately 106 days prior to sacrifice. Exposure
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of parental generation female rats continued through gestation and lactation for a total exposure
period of approximately 118 days. F1 generation rats were exposed to bromodichloromethane in
utero and by consumption of the dam's drinking water during the lactation period. At weaning,
F1 rats (30/sex/concentration) were selected for a postweaning/premating exposure period of at
least 64 days, followed by a cohabitation period of up to 14 days. Exposure continued through
gestation and lactation. F1 generation females delivered litters and the F2 litters were sacrificed
on lactation day 22.
During the course of the experiment, parental and F1 generation rats were evaluated for
viability, clinical signs, water and feed consumption, and body weight. Parental and F1 generation
females were evaluated for estrous cycling (premating and during cohabitation until mating
confirmed and at sacrifice), abortions, premature deliveries, duration of gestation, gestation index,
fertility index, number and sex of offspring per litter, general postpartum condition of dam and
litter, litter size, viability index, lactation index, percent survival, sex ratio, and maternal behavior.
Litters were examined for number and sex of pups, stillbirths, live births, and gross external
alterations. F1 rats selected for continued evaluation were assessed for age at vaginal patency or
preputial separation. At sacrifice, test animals were examined for gross pathology, organ weights,
and histopathology (control and high-dose groups, 10 parental animals/sex; reproductive organs
of 50 and 150 ppm rats suspected of reduced fertility). Male rats were evaluated for sperm
concentration, percent motile sperm, sperm morphology, total number of sperm, and testicular
spermatid counts. Females were evaluated for number and distribution of implantation sites. F1
weanlings not selected for continued evaluation (3 pups/sex/litter, when available) and all F2
weanling rats were evaluated for gross lesions, terminal body weight, and organ weights.
Key findings in the two-generation study reported by CCC (2001c) and Christian et al.
(2002) include the following. The bromodichloromethane dose-equivalent for each drinking
water concentration varied by sex and reproductive status. Average daily doses estimated for the
50, 150 and 450 ppm concentrations were 4.1 to 12.6, 11.6 to 40.2, and 29.5 to 109 mg/kg-day,
respectively, as calculated by the study authors. One death in the 150 ppm group and three deaths
(including one humane sacrifice) in the 450 ppm group were associated with reduced water
consumption, weight loss and/or adverse clinical signs and may have been compound-related.
Adverse clinical signs occurred in parental generation female rats and F1 male and female rats in
the 150 and 450 ppm exposure groups. Compound-related signs included chromorhinorrhea, pale
extremities, urine-stained abdominal fur, and coldness to touch. The study authors attributed
these signs to reduced water consumption. Body weight and body weight gain were significantly
reduced in the 450 ppm parental generation males and females and 150 and 450 ppm F1
generation males and females. The significantly reduced final body weight in 450 ppm parental
generation females was associated with decreased absolute organ weights and increased relative
organ weights when expressed as a percentage of body or brain weight. Absolute and relative
water consumption rates were significantly reduced in parental and F1 generation males and
females at all concentrations of bromodichloromethane. Water intake by parental and F1 animals
was generally reduced by 10 to 20 percent in the 150 and 450 ppm groups when compared to the
controls. Absolute and relative feed consumption rates were reduced in males and females of
both generations at 150 and 450 ppm when compared with the controls. There were no gross
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pathological or histopathological indications of compound-related toxicity. Most indicators of
reproductive or developmental toxicity examined by Christian et al. (2002) were not significantly
affected by bromodichloromethane treatment. However, F1 and F2 generation pup body weights
were reduced in the 150 and 450 ppm groups during the lactation period after the pups began to
drink the water provided to the dams. The F1 generation had statistically significant reductions in
pup body weight at weaning on lactation day 22. Reductions in F2 pup body weight did not reach
statistical significance. Small (<6%), but statistically significant, delays in F1 generation sexual
maturation occurred at 150 (males) and 450 ppm (males and females) as determined by timing of
vaginal patency or preputial separation. The study authors attributed these delays to significant
reductions in body weight at weaning. The values for sexual maturation endpoints in the 150 and
450 ppm exposure groups did not differ significantly from control values when body weight at
weaning was included as a covariate in the analysis. Females rats with vaginal patency not evident
until 40 or 41 days postpartum (i.e., the most delayed) in the 150 and 450 ppm groups had normal
estrus cycles, mated, and produced litters. Estrous cycling in parental generation females was not
affected by exposure to bromodichloromethane. A marginal effect on estrous cyclicity was
observed in F1 females in the 450 ppm exposure group. This effect was reported to be associated
with a higher incidence of rats in the 450 ppm group (5/30) with six or more consecutive days of
diestrus relative to the controls (2/30). The study authors considered this effect to be a secondary
response associated with reduced pup weights and possible inadvertent stimulation of the uterine
cervix during the performance of vaginal smears. Averages for estrous cycles per 21 days,
cohabitation, mating indices, and fertility indices were unaffected by exposure to
bromodichloromethane. Exposure to bromodichloromethane had no effect on anogenital
distances in male or female F2 pups. The results of this study appear to identify NOAEL and
LOAEL values for reproductive effects of 50 ppm (4.1 to 12.6 mg/kg-day) and 150 ppm (11.6 to
40.2 mg/kg-day), respectively, based on delayed sexual maturation. However, the study authors
have questioned whether delayed sexual maturation in F1 males associated with reduced body
weight should be treated as reproductive toxicity or general toxicity, since the root cause appears
to be dehydration brought about by taste aversion to the compound. The parental NOAEL and
LOAEL are also 50 and 150 ppm, respectively, based on reduced body weight and body weight
gain in F0 females and F1 males and females.
2. Dibromochloromethane
Borzelleca and Carchman (1982) evaluated the reproductive toxicity of dibromochloro-
methane in a two-generation study with ICR Swiss mice. The authors used a modified multi-
generation study protocol for this investigation. Groups of 10 males and 30 females (F/0
generation) were exposed to dibromochloromethane in drinking water at concentrations of 0, 0.1,
1.0, or 4.0 mg/mL for seven weeks. The study authors did not estimate average daily doses for
all treated groups. However, they did indicate that the highest drinking water concentration (4.0
mg/mL) corresponded to an average daily dose of 685 mg/kg-day. Using this conversion factor,
drinking water concentrations of 0.1 and 1.0 mg/mL dibromochloromethane were estimated to
correspond to average daily doses of 17 and 171 mg/kg-day, respectively. Following the initial
exposure period, the F/0 mice were mated to produce the F/la litters. Each male mouse was co-
housed for seven days with three randomly selected females. Two weeks after weaning of the
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F/la litters, the F/0 mice were randomly re-mated to produce the F/lb litters. A similar protocol
was followed for the F/lc litters. After a 21-day postnatal period, the F/la and F/lc litters were
sacrificed and necropsied. The F/lb generation was culled. The surviving males and females (10
males and 30 females) were exposed for 11 weeks to dibromochloromethane in drinking water at
concentrations of 0, 0.1, 1.0 or 4.0 mg/mL, and then randomly mated to produce the F/2a and
F/2b generations. A two-week interval occurred between weaning of the F/2a generation and
remating of the F/lb generation to produce the F/2b generation. Thus, parental generations (F0
and F/lb) were exposed continuously to dibromochloromethane in drinking water throughout the
pre-mating, mating, gestation, and lactation periods for a total of 27 and 25 weeks, respectively.
Following weaning of their final litters, both parental generations were sacrificed and necropsied.
The F/2a and F/2b generations were sacrificed and necropsied following a 21-day postpartum
survival period. Additionally, a selected number of pups from the final matings of each generation
(i.e., F/lc and F/2b) were screened for either dominant lethal mutations or teratologic
abnormalities.
Body weight gain and drinking water consumption were recorded weekly and semi-weekly
for the F/0 and F/lb generations, respectively. Mating, gestation, gestation survival, and lactation
survival indices were calculated for each mating. During the 21-day postpartum period, pups
were counted on days 0, 4, 7, 14, and 21, and sexed on days 7, 14, and 21. Viability and lactation
indices were calculated. After sacrifice of all litters except F/lb on day 21, one male and one
female pup per litter were randomly selected for necropsy. For teratology screening, treated
females from the F/0 and F/lb generations were sacrificed on GD 18. The number of
implantations, resorptions, and live and dead fetuses were counted. Fetuses were individually
weighed and examined for gross malformations; a selected subset of fetuses were examined for
either skeletal or visceral anomalies. Statistical analysis was conducted on all endpoints, using
parametric or nonparametric procedures, as appropriate for different endpoints. For statistical
analyses performed on the pups, the sampling unit was the litter. Treatment-related effects were
considered to be statistically significant if the p value< 0.05.
As compared with concurrent controls, final body weights were significantly reduced in
the high-dose males and the mid- and high-dose females of the F/0 and F/lb generations. Water
consumption was unaffected by treatment, indicating that taste aversion was not a factor in the
observed decreases in body weight. Animals in both the F/0 and F/lb generations exhibited
enlarged livers with gross morphological changes, interpreted by the authors as indicative of
hepatotoxicity. The incidence and the severity of these alterations increased with increasing dose,
with 0, 25, 70, and 100% of the F/0 animals and 0, 18, 64, and 100% of the F/lb animal
exhibiting hepatic discoloration, fat accumulation, and/or lesions at 0, 0.1, 1.0, and 4.0 mg/mL,
respectively. Fertility (mating index) was significantly decreased in the high-dose group (4.0
mg/mL) only for the F/2a generation. The gestational index was significantly decreased in the
high-dose group for all three F/l generations. Parental ingestion of 4.0 mg/mL dibromochloro-
methane resulted in (1) decreased litter size in all generations (F/la, F/lb, F/lc, F/2a, and F/2b);
(2) decreased viability index in four of the five generations (F/la, F/lb, F/lc and F/2a); (3)
decreased lactation index in the F/2b generation; and (4) decreased postnatal body weight in the
F/2b generation. Parental ingestion of 1.0 mg/mL dibromochloromethane produced (1) decreased
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litter size in the F/lc generation; (2) decreased viability index in the F/lb generation; (3)
decreased lactation index in the F/lb and F/2b generations; and (4) decreased postnatal body
weight in the F/2b generation. The only statistically significant effect observed at the lowest dose
tested (0.1 mg/mL) was a reduction in postnatal body weight in the F/2b generation on PND 14;
this effect was not noted on PND 7 or 21. No statistically significant increases in dominant lethal
or teratogenic effects were reported in either the F/lc or F/2b generations. Based on decreased
postnatal body weight in the F/2b generation, the marginal LOAEL for
reproductive/developmental toxicity is 17 mg/kg-day and a NOAEL could not be determined.
The developmental LOAEL is considered to be marginal because (1) this effect was only noted in
one of the two litters in the F/2 generation; (2) no other adverse effects were observed at this
dose level; and (3) it was unclear from the report how many litters and pups per litter were
examined for postnatal body weight. For parental toxicity, liver alterations indicative of
hepatotoxicity were clearly evident at the two higher doses in both parental generations. At the
lowest dose tested, hepatic changes were mainly limited to discoloration, presumably due to the
accumulation of fat deposits (Borzelleca and Carchman, 1982); gross morphology was normal,
and histopathologic examination was not conducted. Therefore, the adversity of this effect was
uncertain. Based on these considerations, the lowest dose, 17 mg/kg-day, is considered to be a
marginal LOAEL for parental toxicity (both F0 and Fl/b generations) and a NOAEL could not be
determined.
Ruddick et al. (1983) investigated the reproductive and developmental toxicity of
dibromochloromethane in Sprague-Dawley rats. Pregnant dams (10 to 12 animals per dose
group) were administered gavage doses of 0, 50, 100, or 200 mg/kg-day in corn oil on GD 6-15.
Body weights were measured on GD 1, on GD 1 through GD 15, and before and after caesarean
section on GD 22. On GD 22, females were anesthetized, exsanguinated, and viscera (including
the uteri) were examined for pathological changes. The fetuses were removed, weighed
individually, and examined for viability and external malformations. Two pups per dam were
placed in fixative for histopathological examination. Approximately two-thirds of the remaining
live fetuses were preserved for examination for skeletal abnormalities. The remaining fetuses
were preserved for examination for visceral alterations. Maternal blood was analyzed for
standard hematological and clinical biochemistry parameters. Following gross pathological
examination of the dams, organ weights were collected for liver, heart, brain, spleen, and one
kidney. Tissues from control and high dose dams (5 animals/group) were subject to
histopathological examination. Where chemical related effects were observed, the affected tissues
were also examined in the mid-dose group.
Maternal weight gain was depressed by 25% in the high-dose group relative to controls.
No significant effects on maternal organ weights, hematology and clinical chemistry, number of
resorption sites, number of fetuses per litter, and mean fetal body weight gain were observed in
any of the dose groups. No treatment-related histopathology was noted in either dams or fetuses.
There were no skeletal or visceral fetal anomalies attributed to dibromochloromethane treatment.
Statistical analysis of fetal endpoints was not conducted by the study authors. However,
inspection of the data indicated that there were no dose-related effects (e.g., the number of
affected fetuses/number of affected litters for sternebral aberrations was 3/2, 2/1, 1/1, and 1/1 for
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control, low-, mid-, and high-dose groups, respectively). The power of this experiment was
limited by the small number of litters per dose group. In the absence of observed fetal effects, the
NOAEL for developmental toxicity was 200 mg/kg-day, the highest dose tested, and a LOAEL
could not be determined. Based on significantly decreased maternal body weight gain, the
LOAEL and NOAEL for maternal toxicity were 200 and 100 mg/kg-day, respectively.
NTP (1996) conducted a short-term reproductive and developmental toxicity screen in
Sprague-Dawley rats. Dibromochloromethane was administered in drinking water at
concentrations of 0, 50, 150, or 450 ppm during a study period of 35 days. Males (10/group)
were treated from study days 6 through 34. At study termination, males were submitted for a
thorough examination, which included hematology, clinical chemistry, gross necropsy,
histopathology, and a complete sperm evaluation (count, density, motility, and morphology).
Group A females (10/group) were treated from study days 1 through 34. These females were
mated to treated males on study days 13 through 18 and necropsied on study day 34. Group B
females (13/group) were mated on study day 1 to untreated males, treated from GD 6 through
parturition, and necropsied on postnatal day 5. No hematology, clinical chemistry, or
histopathology was conducted on the females.
Based on measured water consumption, the authors estimated dose levels for the males as
4.2, 12.4, and 28.2 mg/kg-day, for Group A females as 6.3, 17.4, and 46.0 mg/kg-day, and for
Group B females as 7.1, 20.0, and 47.8 mg/kg-day. A few changes in clinical chemistry were
noted for the males. Alkaline phosphatase and 5' nucleotidase were increased at all dose levels in
males, but reached statistical significance only at the low dose for alkaline phosphatase and at the
mid and high dose for 5' nucleotidase. Total serum proteins were also decreased at the high dose
in males. The study authors noted that these changes could reflect mild liver damage. However,
no treatment-related microscopic lesions were observed. No statistically significant effects were
observed on any sperm parameter investigated. No effect was observed on any reproductive or
fertility measure in Group A or B females at any dose. The proportion of male pups was
significantly decreased in Group B females at the high dose compared to the control value. The
study authors did not consider this result to be treatment-related because the control value (0.61)
was unusually high compared to historical values, and the result for the high dose group (0.44)
was within historical background. Based on these data, the authors noted that dibromochloro-
methane was not a reproductive toxicant at doses up to the high dose in either males (28.2 mg/kg-
day) or females (46.0 to 47.8 mg/kg-day). Based on the clinical chemistry changes, the authors
stated that administration of dibromochloromethane may have resulted in general toxicity at all
doses in the male treatment groups. The observed changes in clinical chemistry, however, would
not be considered adverse for the following reasons: absence of clear dose-related response, small
magnitude of the changes, and absence of supporting histopathology data. Therefore, this study
identified NOAEL values of 28.2 mg/kg-day and 47.8 mg/kg-day for males and females,
respectively, for reproductive and systemic effects.
3. Bromoform
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Ruddick et al. (1983) investigated the reproductive and developmental toxicity of
bromoform in Sprague-Dawley rats. Pregnant dams (14 to 15 animals/dose group) were admini-
stered gavage doses of 0, 50, 100, or 200 mg/kg-day in corn oil on GD 6 to 15. Body weights
were measured on GD 1, on GD 6 through GD 15, and before and after pups were delivered by
caesarean section on GD 22. On GD 22, females were sacrificed and body tissues (including the
uterus) were removed for pathological examination. Females were evaluated for the number of
resorption sites and the number of fetuses. Maternal blood samples were collected and evaluated
for standard hematology and clinical chemistry parameters. The liver, heart, brain, spleen, and
one kidney were weighed. Standard histopathology was conducted on control and high dose
females (5/group). All fetuses in all groups were individually weighed, and evaluated for viability
and external malformations. Histopathologic examination was performed on two pups per litter.
Of the remaining live fetuses, approximately two-thirds were examined for skeletal alterations and
one-third for visceral abnormalities.
Maternal weight gain, organ weights, hematology, and clinical chemistry were unaffected
by bromoform treatment. No significant differences between exposed and control groups were
observed for the number of resorption sites, the number of fetuses per litter, fetal weights, fetal
gross malformations, and visceral abnormalities. No treatment-related histopathological effects
were noted in either the dams or fetuses. However an elevation in the incidence of skeletal
anomalies, including the presence of a 14th rib, wavy ribs, and interparietal bone deviations was
reported in treated animals. An increase in wavy ribs was only observed in the high dose group.
The number of affected fetuses /number of affected litters for the presence of a 14th rib was 3/3,
4/3, 4/3 and 7/5 in the 0, 50, 100, and 200 mg/kg-day groups, respectively. The incidence of
sternebral aberrations (number of affected fetuses/number of affected litters) was 1/1, 5/3, 6/5,
13/8 in the 0, 50, 100, and 200 mg/kg-day groups, respectively. Statistical significance for
fetotoxic endpoints was not reported by the study authors. A statistical analysis (using the Fisher
Exact test) was conducted on the published data and demonstrated that the increase in sternebral
anomalies was significantly different from controls at the highest dose tested (200 mg/kg-day). A
trend test showed a statistically significant dose-related trend (p< 0.002) for this endpoint;
stepwise analysis indicated that the trend was no longer significant when the two highest doses
(i.e., 200 and 100 mg/kg-day) were omitted from the analysis. These findings suggest that the
LOAEL and NOAEL for developmental toxicity were 100 and 50 mg/kg-day, respectively. In the
absence of observed maternal effects, the NOAEL for maternal toxicity was 200 mg/kg-day, and a
LOAEL could not be determined.
NTP (1989b) investigated the effect of bromoform on fertility and reproduction in Swiss
CD-I mice using a continuous breeding protocol. Twenty male-female pairs were administered
daily doses of 50, 100, or 200 mg/kg-day by gavage in corn oil and forty male-female pairs were
dosed with the corn oil vehicle only. Dose selection was based on a 14-day range-finding study.
The 105-day dosing period included a seven-day precohabitation phase and a 98-day cohabitation
phase. The parameters evaluated for this study were fertility, litters per pair, live pups per litter,
proportion of pups born alive, sex of live pups, or pup body weights. The last litter born
(generally the fifth litter) in the control and 200 mg/kg-day groups during a holding period
following the continuous breeding phase were reared by the dams, weaned and raised to sexual
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maturity (approximately 74 days) while receiving the same treatment (vehicle control of 200
mg/kg-day bromoform) as their parents. At sexual maturity, males and females from different
litters within the same treatment group were cohabited for seven days and then housed
individually until delivery. The endpoints for this mating trial were the same as for the parental
generation. At study termination, the Fx mice were weighed, necropsied and evaluated for
selected organ weights, epididymal sperm motility, and sperm morphology. Selected organs were
fixed for histopathological examination.
In the 200 mg/kg-day treatment group, the body weights of dams at delivery were
consistently less than the control group value. The reduction in body weight was statistically
significant after delivery of the first, second, fourth, and fifth litters. The fertility index for the
parental generation was 100% for the control and treated groups (a breeding pair was designated
as fertile if they produced at least one live or dead pup). There was no detectable effect of
treatment on the number of litters per pair, the number of live pups per litter, the proportion of
pups born alive, the sex of live pups, or pup body weights. The gestational period was similar
across groups. However, postnatal survival of Fx pups in the 200 mg/kg-day group was
significantly lower than in the control group. The study authors reported that this difference was
primarily attributable to three dams who lost all of their pups by postnatal day 4. One dam in the
control group also lost her litter by postnatal day 4. The study authors noted that this result is
consistent with a treatment effect on early maternal behavior, early lactational failure, and/or the
postnatal developmental processes. When F, mice were cohabited for one week, no effect of
treatment on mating index or fertility was observed. There were no significant differences relative
to control values for the number of live pups per litter (male, female, or combined), the proportion
of live pups, the proportion of male pups, or pup weight at birth. At sacrifice, male and female Fx
mice administered 200 mg/kg-day exhibited increased relative liver weights and decreased relative
kidney weights as compared with control values. The mean body weight for F, males was
significantly less than the mean weight of the male control group. Histopathological evaluation
revealed minimal to moderate hepatocellular degeneration in the livers of high-dose F, male and
female mice. Bromoform treatment had no effect on epididymal sperm density, motility, or
morphology in F, males. No treatment-related histologic lesions were observed in the seminal
vesicles, coagulating glands, or prostate glands of males, or in the lung, kidney, or thyroid gland
of males or females. Based on liver histopathology, decreased postnatal survival, and other signs
of toxicity (e.g., increased relative liver and decreased relative kidney weights) in F, mice of both
sexes at the highest dose tested, the LOAEL for developmental toxicity is 200 mg/kg-day, and the
NOAEL is 100 mg/kg-day. Based on consistently decreased body weights of pregnant dams at
delivery, the LOAEL for maternal toxicity is 200 mg/kg-day and the NOAEL is 100 mg/kg-day.
Table V-9 Summary of Reproductive Studies of Brominated Trihalomethanes
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromodichloromethane
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Table V-9 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Ruddick et al.
(1983)
Rat
Sprague-
Dawley
Gavage
(Corn oil)
F
9-14
GD 6-15
0
50
100 (NOAEL)1
200 (LOAEL)
Developmental toxicity study.
Statistically decreased maternal wt.
gain at high dose (38%); increased
litter incidence of sternebral
aberrations. Statistical significance not
evaluated for fetotoxic endpoints by
study authors; statistical analysis
conducted on published data for fetal
effects. Trend test indicated statistical
effect for sternebral anomalies at
highest dose tested. Maternal LOAEL
and NOAEL are 200 and 100 mg/kg-
day, respectively.
Klinefelter et al.
(1995)
Rat
F344
Drinking
Water
M
7
52 weeks
0
22 (NOAEL)
39 (LOAEL)
Reproductive toxicity study.
Decreased sperm velocity at 39 mg/kg-
day. No histopathological alterations
noted in any reproductive tissue
examined.
Narotsky et al.
(1997)
Rat
F344
Gavage
(Corn oil)
F
12-14
GD 6-15
0
25 (NOAEL)
50 (LOAEL)
75
Developmental toxicity study
comparing the use of different gavage
dosing vehicles. Reduced maternal
weight gain GD 6-8 and lacrimation at
50 and 75 mg/kg-day. Full litter
resorption (FLR) observed at 50 and
75 mg/kg-day (8% and 83%,
respectively). No effects on postnatal
survival, pup weight, or pup survival
in surviving litters. ED05 and BMD
for FLR calculated by study authors as
33.3 and 11.3 mg/kg-day,
respectively. Maternal LOAEL and
NOAEL are 50 and 25 mg/kg-day,
respectively.
Narotsky et al.
(1997)
Rat
F344
Gavage
(Aqueous)
F
12-14
GD 6-15
0
25 (NOAEL)
50 (LOAEL)
75
Developmental toxicity study
comparing the use of different gavage
dosing vehicles. Reduced maternal
weight gain GD 6-8 at all dose levels.
Full litter resorption (FLR) observed
at 50 and 75 mg/kg-day (17 and 21%,
respectively). No effects observed on
postnatal survival, pup weight, or pup
survival in surviving litters. Maternal
LOAEL is 25 mg/kg-day; maternal
NOAEL not determined.
NTP (1998)
Rat
Sprague-
Dawley
Drinking
Water
M
5-10
25-30 days
0
8
41
68 (NOAEL)
Reproductive/developmental toxicity
study. Decreased food and water
consumption; decreased body weight.
No dose-related changes in
reproductive/developmental
parameters
NTP (1998)
Rat
Sprague-
Dawley
Drinking
Water
F
5-10
25-30 days
0
14
72
116 (NOAEL)
Reproductive/developmental toxicity
study. Decreased food and water
consumption; decreased body weight.
No dose-related changes in
reproductive/developmental
parameters
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Table V-9 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bielmeier et al.
(2001)
Rat
F344
Gavage
(Aqueous)
F
8-13
GD 6-15
GD 6-10
GD 11-15
0
75 (LOAEL)
Critical exposure period study. Full
litter resorption observed in animals
treated on GD 6-10, but not in animals
treated on GD 11-15.
Bielmeier et al.
(2001)
Rat
F344/
Sprague-
Dawley
Gavage
(Aqueous)
F
12-14
GD 6-10
F344
0
75 (LOAEL)
Sprague-Dawlev
0
75
100 (NOAEL)
Strain comparison study. Full litter
resorption observed in F344 rats but
not in concurrently dosed Sprague-
Dawley rats.
Bielmeier et al.
(2001)
Rat
F344
Gavage
(Aqueous)
F
8-11
GD 9
0
75 (LOAEL)
100
Hormone profile study. Full litter
resorption observed at both doses.
Decreased serum progesterone levels
in F344 rats which experienced FLR.
CCC (2000c)
Christian et al.
(2001b)
Rat
Sprague-
Dawley
Drinking
Water
F, M
10
64-74 days
Males2
(days 1-64
i.e., for 14
days
premating,
during
mating,
and for -
6 weeks
following
mating)
Females2
(days 1-74,
i.e., for 14
days
premating,
during
mating,
GD 0-21,
lactation
days 1-29)
0 ppm
50 ppm (NOAEL)
150 ppm
(LOAEL)
450 ppm
1350 ppm
Range finding
reproductive/developmental toxicity
study. Decreased body weight gain
and terminal body weight (>10%) in
males at highest dost tested but no
apparent effects on reproductive
endpoints at any dose.
Maternal toxicity (reduced body
weight and body weight gain and
decreased food and water
consumption) at 150 ppm and higher.
Dose-dependent decreases in mean pup
weight gain and pup weights beginning
on lactation day 5-29 in 3 highest
dose groups. Decreased pup body
weight gain and body weight also
observed in 3 highest dose groups in
pups treated for one week postweaning
at parental drinking water
concentrations.
Reproductive/developmental LOAEL
and NOAEL are 150 and 50 ppm,
respectively, based on decreased pup
wt and wt gain; parental LOAEL and
NOAEL are 150 and 50 ppm,
respectively, based on reduced body wt
gain and wt. Findings confounded by
effects of decreased water consumption
at various time points during
treatment.
CCC (2000d)
Christian et al.
(2001a)
Rat
Sprague-
Dawley
Drinking
Water
F
25
GD 6-21
0
2.2
18.4
45.0 (NOAEL)
82.0 (LOAEL)
Developmental toxicity study.
Decreased maternal body weight and
body weight gain at 45.0 mg/kg-day.
Developmental LOAEL based on
slightly decreased number of
ossification sites in the hindlimb
(metatarsals and phalanges) and
forelimb (phalanges). Maternal
LOAEL and NOAEL are 82.0 and
45.0 mg/kg-day, respectively.
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Table V-9 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
CCC (2000a)
Christian et al.
(2001b)
Rabbit
New
Zealand
White
Drinking
Water
F
5
GD 6-29
0.0
4.9
13.9
32.3
76.3 (NOAEL)
Range finding developmental study.
Decreased maternal body weight gain
and water and feed consumption at all
tested doses. No treatment-related
changes in reproductive or
developmental endpoints. Study
authors considered maternal LOAEL
to be < 4.9 mg/kg-day, based on
significantly reduced body weight
gain.
CCC (2000b)
Christian et al.
(2001a)
Rabbit
New
Zealand
White
Drinking
Water
F
25
GD 6-29
0
1.4
13.4
35.6
55.3 (NOAEL)
Developmental toxicity study.
Reduced maternal weight gain at 35.3
mg/kg-day. No dose-related changes
in reproductive or developmental
parameters. Maternal LOAEL and
NOAEL are 35.3 and 13.4 mg/kg-day,
respectively.
CCC (2002)
Rat
Sprague-
Dawley
Drinking
Water
M,F
30
Up to 118
days
Males
F0: 106 d (incl. 70
d pre-mating)
¥{. 64 d post-
weaning, 14 d
cohab.
Females
F0: 118 d (incl.
gest., lactation
F: 64 d post-
weaning, 14 d
cohab., gest.,
lactation
0
50 (NOAEL)
150 (LOAEL)
450
Reproductive/developmental LOAEL
and NOAEL are 150 and 50 ppm,
respectively, based on delayed sexual
maturation in ¥1 males; parental
LOAEL and NOAEL are 150 and 50
ppm, respectively, based on reduced
body wt gain and wt in F0 females and
Fj males and females. Findings
confounded by effects of decreased
water consumption as a result of taste
aversion to the test compound.
Dibromochloromethane
Borzelleca and
Carchman
(1982)
Mouse
ICR Swiss
Drinking
Water
M
F
10
30
25-27
weeks
0
17 (marginal
LOAEL)
171
685
Multi-generation reproductive toxicity
study. Significant high-dose effects
include decreased gestational index in
F1 generation at high dose and
decreased litter size in F1 and F2
generations. Significant mid-dose
effects include decreased litter size,
decreased viability index, decreased
lactation index, and decreased
postnatal body weight in some F1
and/or F2 generations. Only
significant low-dose effect is reduced
postnatal body wt in F/2b generation
on postnatal day 14. Hepatic effects
observed in both parental generations
at all doses; liver effects marginal at
low dose. Parental marginal LOAEL
is 17 mg/kg-day; parental NOAEL
not determined.
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Table V-9 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Ruddick et al.
(1983)
Rat
Sprague-
Dawley
Gavage
(Corn oil)
F
10-12
GD 6-15
0
50
100
200 (NOAEL)
Developmental toxicity study.
Significantly depressed maternal wt.
gain at high dose (25%); increased
maternal relative liver wt. and kidney
Statistical significance of fetal
endpoints not evaluated by study
authors. Based on data inspection, no
dose-related skeletal or visceral effects
observed in litters. Maternal LOAEL
and NOAEL are 200 and 100 mg/kg-
day, respectively.
NTP (1996)
Rat
Sprague-
Dawley
Drinking
Water
M
10
29 days
4.2
12.4
28.2 (NOAEL)
Reproductive/developmental toxicity
study. No treatment-related effects on
measured sperm parameters
NTP (1996)
Rat
Sprague-
Dawley
Drinking
Water
F
10
35 days
6.3
17.4
46.0 (NOAEL)
Reproductive/developmental toxicity
study. Exposure occurred during a 6-
day mating period and most/all of
gestation. No clearly adverse effect on
any reproductive or developmental
endpoint at tested doses
NTP (1996)
Rat
Sprague-
Dawley
Drinking
Water
F
13
-16 days
(GD 6 to
parturition)
7.1
20.0
47.8 (NOAEL)
Reproductive/developmental toxicity
study. No clearly adverse effect on
any reproductive or developmental
endpoint at tested doses
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Table V-9 (cont.)
Reference
Species
Route
Sex
Number
per dose
group
Duration
Dose
(mg/kg-day)
Results
Bromoform
Ruddick et al.
Rat
Gavage
F
14-15
GD 6-15
0
Developmental toxicity study. No
(1983)
Sprague-
Dawley
(Corn oil)
50 (NOAEL)
100 (LOAEL)
200
statistically significant maternal
effects. Apparent treatment-related
increases in sternebral aberrations and
other skeletal endpoints. Statistical
significance of fetotoxic endpoints not
evaluated by study authors. Statistical
analysis conducted on published data
indicated significant increase in
sternebral aberrations at two highest
doses. Maternal NOAEL is 200
mg/kg-day; LOAEL not determined.
NTP (1989b)
Mouse
Gavage
M
20
105 days
0
Continuous breeding reproductive
ICR Swiss
(oil)
F
20
50
100 (NOAEL)
200 (LOAEL)
toxicity protocol. Maternal body
weights significantly decreased at
highest dose tested. Decreased
postnatal survival, organ wt. changes,
and liver histopathology observed in
F1 mice of both sexes in high-dose
group. No noted effects on fertility,
litters/pair, live pups/litter; proportion
of live births, sex of live pups, or pup
body weight. Maternal LOAEL and
NOAEL are 200 and 100 mg/kg-day,
respectively.
1 NOAEL and LOAEL values reported in this column are for developmental/reproductive toxicity effects. The
NOAEL and LOAEL values for parental toxicity are reported in the "Results" column.
2 Doses for this study are presented as ppm in drinking water; due to marked changes in adult female water
consumption during different physiologic stages (i.e., pre-mating, mating, gestation, and lactation), it is not
possible to convert administered drinking water concentrations into biologically meaningful average daily
doses.
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F. Mutagenicity and Genotoxicitv
1. Bromodichloromethane
The results of in vivo and in vitro tests conducted to evaluate the mutagenicity, genotoxi-
city, and neoplastic transformation potential of bromodichloromethane are summarized in Table
V-10 at the end of this section.
In Vitro Assays
Simmon and Tardiff (1978) reported that nonactivated bromodichloromethane was
mutagenic in S. typhimurium strain TA100 when assayed in a desiccator containing the test
compound in the atmosphere. The minimum amount of bromodichloromethane required to elicit
a mutagenic response following addition to the desiccator was 600 |imol.
Ishidate et al. (1982) assayed the mutagenicity of bromodichloromethane in S.
typhimurium strain TA100 in the presence and absence of rat liver S9 fraction. Increased
mutation frequencies were observed only in the absence of S9 activation. In contrast,
chromosomal aberrations in Chinese hamster fibroblasts were observed in the presence, but not
the absence, of S9 fraction. The concentrations tested in these assays were not reported.
Nestmann and Lee (1985) investigated the mutagenicity of bromodichloromethane at 12
to 1,200 |iM in S. cerevisiae strains D7 and XV185-14C in the presence or absence of S9
activation. No clear increase in convertants or in revertants of strain XV185-14C was observed
for bromodichloromethane in the presence or absence of S9 activation.
NTP (1987) reported that bromodichloromethane was not mutagenic when tested using a
preincubation protocol in S. typhimurium strains TA1535, TA1537, TA98, or TA100 at
concentrations reaching cytotoxic levels (20 |imol/plate; 3,333 |ig/plate). Testing was done in the
absence of S9 and in the presence of S9 prepared from Aroclor-induced male hamster or rat liver.
NTP concluded that the negative results may have been due to volatilization of the test compound
from the test system. Bromodichloromethane was not mutagenic in the mouse lymphoma
L5178Y/TK+/" assay in the absence of S9, but did induce dose-related increases in forward
mutations at S9-activated concentrations greater than or equal to 2,000 |iM (300 |ig/mL).
Cytogenetic tests with Chinese hamster ovary cells (CHO) cells were reported in this study and
also by Anderson et al. (1990). There was no evidence of induction of chromosomal aberrations
following treatment with up to 30,000 |iM (5,000 |ig/mL) in either the presence or absence of
exogenous metabolic activation. There was also no evidence of sister chromatid exchanges
induced by the nonactivated material. In the presence of S9 activation, one of three assays
resulted in a positive response at doses greater than or equal to 24,400 |iM (4,000 |ig/mL).
These results are difficult to evaluate because cytotoxicity was observed at similar concentrations
in the other trials.
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Varma et al. (1988) tested bromodichloromethane for mutagenicity in S. typhimurium
strains TA1535, TA1537, TA98, and TA100. In the absence of S9 fraction, bromodichloro-
methane at nonactivated concentrations of 2.4 to 3.2 |imol/plate induced mutations in strain
TA1537. There was no effect in the other strains.
Bromodichloromethane was positive for the induction of DNA damage in the presence
and absence of exogenous activation, based on the results of the SOS chromotest (LeCurieux et
al., 1995). Bromodichloromethane gave a negative result in the fluctuation test modification of
the S. typhimurium reverse mutation assay.
Several studies have evaluated induction of sister chromatid exchanges following exposure
to bromodichloromethane. Morimoto and Koizumi (1983) investigated the ability of
bromodichloromethane to induce sister chromatid exchanges in human lymphocytes in vitro in the
absence of S9 activation. Bromodichloromethane caused a dose-dependent increase in sister
chromatid exchanges. The increased incidence was significant (p < 0.05) at concentrations greater
than or equal to 400 |iM. The potential of S9-activated bromodichloromethane to induce sister
chromatid exchanges in vitro was also investigated by Sobti (1984). A dose of 100 |iM increased
the frequency of sister chromatid exchange in rat liver cells. In human lymphocytes, a 100-fold
greater concentration of bromodichloromethane was required to elicit the same effect on sister
chromatid exchange when compared to dibromochloromethane (100 |iM vs. 1 |iM.). Fujie et al.
(1993) observed a statistically significant, dose-related increase in sister chromatid exchange in rat
erythroblastic leukemia K3D cells treated with bromodichloromethane in the absence of
exogenous activation. Bromodichloromethane also appeared to give a positive response in the
presence of exogenous metabolic activation, although the study protocol and results with negative
controls were described less clearly for this phase of testing.
Bromodichloromethane was tested in the mouse lymphoma assay as part of an
international collaborative program under the auspices of the Japanese Ministry of Health and
Welfare (Sofuni et al., 1996). The results of this assay were equivocal. One laboratory obtained
a positive result in the activated phase, but this result was not confirmed by a second laboratory.
Results in the nonactivated phase were negative or equivocal due to poor viability of the solvent
control cell cultures.
Matsuoka et al. (1996) conducted a chromosome aberration assay with Chinese hamster
lung fibroblast (CHL/IU) cells exposed to bromodichloromethane in tightly capped flasks. A
weak induction of chromosome aberrations was observed for bromodichloromethane in the
presence and absence of exogenous metabolic activation.
Several studies have investigated the mutagenicity of bromodichloromethane in strains of
Salmonella typhimurium engineered to express the rat theta-class glutathione S-transferase Tl-1
gene (GSTT1-1). These studies provide evidence for a third mechanism of brominated trihalo-
methane activation and thus are discussed in detail. Pegram et al. (1997) utilized two new strains
of TA1535-derived Salmonella to investigate glutathione S-transferase-mediated bioactivation of
bromodichloromethane. One strain had been transfected with the GSTT1-1 gene (+GST) and the
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other strain had the same gene inserted in a non-functioning orientation (-GST). These strains
were used in base-substitution revertant colony assays following 24 hour exposures to
concentrations of bromodichloromethane ranging from 200 to 4,800 ppm. The agar
concentration resulting from a 24 hour exposure to 4,800 ppm bromodichloromethane was 0.67
mM. Bromodichloromethane increased the number of revertant colonies in each strain of
Salmonella tested (+GST, -GST and TA1535). The frequency of the revertants in TA1535 was
significantly increased above the spontaneous level at the three highest concentrations tested
(highest concentration 4,800 ppm; intermediate concentrations not explicitly stated), while
frequency was increased in the -GST strain only at the highest concentration. In contrast, there
was a dramatic, dose-dependent increase in bromodichloromethane-induced mutations in the
+GST strain when compared to the -GST control strain (an 18-fold increase at the 4800 ppm
bromodichloromethane concentration). When chloroform was tested for comparative purposes, a
positive response was observed only at the two highest concentrations tested (19,200 and 25,600
ppm). These results provide evidence that the mutagenicity of bromodichloromethane is
enhanced by GST-mediated conjugation with GSH. The comparatively low affinity of the GST-
mediated pathway for chloroform suggests that different trihalomethanes can induce mutations by
different mechanisms.
DeMarini et al. (1997) further investigated the role of glutathione S-transferase activity in
mediating the mutagenicity of bromodichloromethane in Salmonella typhimurium. Strains of
Salmonella utilized in this investigation included RSJ100, which has been engineered to express
the GSTT1-1 gene and TPT100, which has the GSTT1-1 gene inserted in a non-functioning
orientation. Mutagenicity was assayed using a Tedlar bag vaporization technique.
Bromodichloromethane (3,200 ppm) induced a 44-fold increase in revertant colonies in the
RSJ100 strain of Salmonella when compared to background revertant frequency. The spectrum
of bromodichloromethane-induced mutations at the hisG46 allele in strain RSJ100 was analyzed
using the colony probe hybridization method. This analysis revealed that 99% of the mutations
were GC—>AT. A non-brominated halomethane, dichloromethane, was used in S. typhimurium
strain TA100 (which does not contain the GSST1-1 expressing of) for comparison. In contrast to
bromodichloromethane-induced mutations in RSJ100, only 15% of the mutations induced by
dichloromethane in the non-GST-expressing strain TA100 were GC—>AT type mutations. This
result suggests that over-expression of GSTT1-1 in strain RSJ100 enhanced the mutagenicity of
bromodichloromethane and induced a specific type of mutational lesion in Salmonella. The
mutagenicity of dibromochloromethane and bromoform was also markedly enhanced in the GST-
expressing strain (discussed below), suggesting that the brominated trihalomethanes are bio-
activated by a similar pathway. In contrast, the mutagenicity of chloroform was not enhanced,
indicating that chloroform and the brominated trihalomethanes may be activated via different
mechanisms. Proposed routes for GST-mediated bioactivation are illustrated in Figure V-2.
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Figure V-2 Proposed Routes for GST-Mediated Metabolic Activation of Trihalomethanes
[IV] GB-CHrCNpJdG
A
CH2X2 ¦
A
(+ 2-dG)
*asn
X OH
| 4H20 I
GS—CH2 > GS—CH2
(41>0 k ... (41X>
[¦1
CHX3 ¦>
: x
GS—CH [V]
i i
h2c=o pq
(-gkh>
NAD+/FDH
(+ 2"-dG>
IVI] GS-CH-( lSp)dG
X
GS-CH
[Ilia]
(-GSH)
+h2o
(-gsh>
0
u
HC-OH
[¦¦lb]
A t
[VII] CHO-( N2)dG,dC
Note: Solid arrows represent known pathways (modified from Gargas et al. (1986)); dashed arrows represent
proposed pathways lacking direct experimental evidence. (I), .S'-( l-halomethyl)GSH: (II), formaldehyde; (Ilia), S-
(l-formyl)GSH; (Illb), formic acid; (IV), .S'-| 1 -(.Y--dco.\yguanosinyl)mcthvl |GSH adduct (Thier et al., 1993); (V),
>S-( 1.1 -dihalomethvl)GSH; (VI), 5'-[l-halo(7V2-deoxyguanosinyl)methyl]GSH adduct; and (VII), TV-formyl adduct on
either G or C. Adapted from DeMarini et al. (1997).
In Vivo Assays
Ishidate et al. (1982) investigated the in vivo clastogenicity of bromodichloromethane in
ddY mice, MS mice, and Wistar rats. Doses of 125 to 500 mg/kg-day were administered in olive
oil by intraperitoneal injection, and the animals were sacrificed at 18, 24, 30, 48, and 72 hours
after dosing. No significant induction of micronucleus formation in bone marrow was observed in
either mice or rats.
Morimoto and Koizumi (1983) investigated the potential of bromodichloromethane to
induce sister chromatid exchanges in male ICR/SJ mice. Animals were given doses of 0, 25, 50,
100, or 200 mg/kg-day for four days by olive oil gavage. Bromodichloromethane produced a
roughly linear dose-dependent increase in sister chromatid exchange frequency. The increase was
statistically significant (p < 0.05) at 50 mg/kg-day. The authors noted that the concentrations
required to produce an increased incidence of sister chromatid exchange were on the order of
1,000 to 10,000 times higher than the concentrations typically found in drinking water.
Hayashi et al. (1988) measured induction of micronucleated polychromatic erythrocytes in
ddY mice by intraperitoneal administration of bromodichloromethane at single doses up to 500
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mg/kg in corn oil. No evidence of clastogenicity was observed. There was no clear evidence of
toxicity or cytotoxicity in the target tissue.
Fujie et al. (1990) analyzed chromosome aberrations in bone marrow from Long-Evans
rats (3/sex/dose) following oral (males only) or intraperitoneal (males and females) exposure to
bromodichloromethane. Oral administration was by gavage in saline for five consecutive days,
and the animals were sacrificed 18 hours after the last dose. Bromodichloromethane induced
dose-related increases in chromatid and chromosome breaks. A more pronounced increase in
clastogenic activity was observed following a single intraperitoneal dose, with significant (p <
0.05) effects at 16.4 mg/kg.
Hayashi et al. (1992) evaluated induction of micronuclei in mouse peripheral blood
erythrocytes by bromodichloromethane. Groups of four male ddY mice received an intraperi-
toneal injection of 0, 25, 50, or 100 mg/kg bromodichloromethane in physiological saline once a
week for 5 weeks. Micronuclei were evaluated 1 week after the last dose. No evidence of
micronucleus induction was observed. One low-dose mouse died, and weight loss was observed
in all treatment groups during exposure.
Potter et al. (1996) investigated the effect of bromodichloromethane on DNA strand
breakage in male F344 rats. Test animals received 0.75 or 1.5 mmol/kg of bromodichloromethane
in 4% Emulphor® by gavage for 1, 3, or 7 days. The administered doses corresponded to 123 or
246 mg/kg-day. One day after administration of a single dose, DNA strand breaks in the kidney
were analyzed using the alkaline unwinding procedure. No treatment-related effect was observed
at either dose level.
Stocker et al. (1997) investigated the in vivo genotoxicity of bromodichloromethane in an
unscheduled DNA synthesis assay in the livers of bromodichloromethane treated rats. Male
Sprague-Dawley rats (4 animals per group) were administered a single dose of 0 (control), 135
or 450 mg/kg bromodichloromethane via gavage in aqueous 1% methylcellulose. These doses
were selected by the authors to correspond to 30% and 100% of the calculated maximum
tolerated dose (MTD) for this compound. Analysis of hepatocytes for unscheduled DNA
synthesis was conducted 2 and 14 hours after treatment. There was no evidence of increased
DNA synthesis in hepatocytes at any tested dose of bromodichloromethane.
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Table V-10 Summary of Mutagenicity, Genotoxicity, and Neoplastic Transformation Data
for Bromodichloromethane
Endpoint
Assay System
Results
(with/without
activation)1*
References
In Vitro Studies
Gene mutation
Salmonella typhimurium
TA100a
NT/+
Simmon and Tardiff (1978)
TA100b
-/+
Ishidate et al. (1982)
TA98, TA100, TA1535,
TA1537b
-/-
NTP (1987)
TA1537
TA1535, TA98, TA100b
-/+
-/-
Varma et al. (1988)
RSJ100
NT/+
DeMarini et al. (1997)
TA1535, +GST, -GST
NT/+
Pegram et al. (1997)
Mouse lymphoma cellsb
+/-
NTP (1987)
Mouse lymphoma cells
+c/-c
Sofuni et al. (1996)
Chromosome aberration
Chinese hamster fibroblastsb
+/-
Ishidate et al. (1982)
(Chinese hamster ovary cells
-/-
NTP (1987); Anderson et al.
(1990)
Chinese hamster lung
fibroblasts"
+/+
(weak)
Matsuoka et al. (1996)
DNA damage
Saccharomyces cerevisiae a
-/-
Nestmann and Lee (1985)
SOS chromotest
+/+
LeCurieux et al. (1995)
Sister chromatid
exchange
Human lymphocytes"
NT/+
Morimoto and Koizumi (1983)
Human lymphocytesa
+/NT
Sobti (1984)
Rat liver cells a
+/NT
Sobti (1984)
[Chinese hamster ovary cells
-7-
NTP (1987); Anderson et al.
(1990)
In Vivo Studies
Micronuclei
Mouse bone marrow cells
and rat cells
-
Ishidate et al. (1982)
Mouse bone marrow cells
-
Hayashi et al. (1988)
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Table V-10 (cont.)
Endpoint
Assay System
Results
(with/without
activation)1*
References
Mouse peripheral blood
erythrocytes (ip)
-
Hayashi et al. (1992)
Chromosome
aberrations
Rat bone marrow cells (oral)
+
Fujie et al. (1990)
Rat bone marrow cells (ip)
+
Fujie et al. (1990)
Rat kidney cells
-
Potter et al. (1996)
Unscheduled DNA
synthesis
Rat liver cells
-
Stacker et al. (1997)
Sister chromatid
exchange
Mouse bone marrow cells
+
Morimoto and Koizumi (1983)
NT = Not Tested
a Assay was conducted in a closed system.
b Authors did not specify whether or not the assay was conducted in a closed system.
c Equivocal results were obtained.
d With/without activation applies to in vitro tests only.
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2. Dibromochloromethane
The results of in vivo and in vitro tests conducted to evaluate the mutagenicity,
genotoxicity, and neoplastic transformation potential of dibromochloromethane are summarized in
Table V-l 1 at the end of this section.
In Vitro Assays
Simmon and Tardiff (1978) reported that nonactivated dibromochloromethane was
mutagenic in S. typhimurium strain TA100 when assayed in a desiccator containing the test
compound in the atmosphere. The minimum amount of dibromochloromethane required to elicit
a mutagenic response following addition to the desiccator was 57 |imol.
Ishidate et al. (1982) assayed the mutagenicity of dibromochloromethane in S.
typhimurium strain TA100 in the presence and absence of rat liver S9 fraction. Increased
mutation frequencies were observed only in the absence of S9 activation. In contrast,
chromosomal aberrations in Chinese hamster fibroblasts were observed in the presence, but not
the absence, of S9 fraction. The concentrations tested in these assays were not reported.
NTP (1985) reported that dibromochloromethane (0.5 to 50 |imol/plate; 100 to 10,000
|ig/plate) was not mutagenic in strains TA1535, TA1537, TA98, or TA100 when tested in the
presence or absence of Aroclor-induced Sprague-Dawley rat or Syrian hamster liver S9 fractions.
Volatilization of the test compound was proposed as a possible explanation for the negative
results.
Nestmann and Lee (1985) investigated the mutagenicity of dibromochloromethane at
concentrations of 11 to 5,700 |iM in S. cerevisiae strains D7 and XV185-14C in the presence or
absence of S9 activation. No clear increase in convertants or in revertants of strain XV185-14C
were observed in the presence of S9-activated dibromochloromethane. In the absence of S9
activation, an increased incidence of gene convertants in strain D7 was observed at concentrations
greater than 1,140 |iM. There was no effect on the incidence of revertants under the same
conditions. The high dose of dibromochloromethane was cytotoxic.
Varma et al. (1988) tested dibromochloromethane for mutagenicity in S. typhimurium
strains TA1535, TA1537, TA98, and TA100. Dibromochloromethane produced a significantly
increased mutation frequency at the lowest S9-activated concentration (0.12 (amol/plate) in all
four strains. Dibromochloromethane at the same concentration also resulted in increased
mutation frequencies in strains TA1535 and TA1537 in the absence of S9 fraction. Higher
concentrations had no clear effect on mutation frequency. This spike in mutation frequency at the
low dose with similar responses in strains that detect frameshifts and those that detect base
substitutions is very unusual. It is possible that the reported data may have resulted from
cytotoxicity, although the number of revertants at the nonmutagenic doses was comparable to
background levels.
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Dibromochloromethane induced mutations at the tk locus of L5178Y mouse lymphoma
cells when tested at concentrations greater than or equal to 480 |iM in screw-capped tubes. The
material was tested only in the absence of S9 activation (McGregor et al., 1991).
Loveday et al. (1990) found that dibromochloromethane did not induce chromosome
aberrations in CHO cells with S9-activation at concentrations that caused cell-cycle delay
(12,200 |iM) or in the absence of S9-activation at concentrations that were cytotoxic with a
standard harvest time (6,000 |iM). Sister chromatid exchange was induced in CHO cells by S9-
activated dibromochloromethane at 3,600 |iM with a delayed cell harvest, while the nonactivated
test material had no effect at concentrations up to cytotoxic levels (1,200 |iM; 247 |ig/mL).
Morimoto and Koizumi (1983) investigated the ability of dibromochloromethane to induce
sister chromatid exchanges (SCE) in human lymphocytes in vitro in the absence of S9 activation.
Addition of dibromochloromethane resulted in a dose-dependent increase in SCE. The increased
incidence was significant (p < 0.05) at concentrations greater than or equal to 400 |iM.
The potential of S9-activated dibromochloromethane to induce sister chromatid exchanges
in vitro was also investigated by Sobti (1984). A dose of 100 |iM produced an increased
frequency of sister chromatid exchange in rat liver cells. In human lymphocytes, 1 |iM
dibromochloromethane produced the same effect as 100 |iM bromodichloromethane.
Fujie et al. (1993) observed a statistically significant, dose-related increase in sister
chromatid exchange in rat erythroblastic leukemia K3D cells treated with dibromochloromethane
in the absence of exogenous activation. Dibromochloromethane had the weakest response among
the brominated trihalomethanes tested. Dibromochloromethane also appeared to give a positive
response in the presence of exogenous metabolic activation, although the study protocol and
results with negative controls were less clear for this phase of testing.
LeCurieux et al. (1995) evaluated the induction of DNA damage by
dibromochloromethane in the presence and absence of exogenous activation using the SOS
chromotest. Dibromochloromethane exposure gave a positive result for induction.
Dibromochloromethane gave negative results in the fluctuation test modification of the S.
typhimurium reverse mutation assay.
Matsuoka et al. (1996) conducted a chromosome aberration assay with Chinese hamster
lung fibroblast (CHL/IU) cells exposed to dibromochloromethane in tightly capped flasks.
Dibromochloromethane induced polyploidy in the absence of S9 fraction, but not in the presence
of S9. The study authors considered activated dibromochloromethane marginally positive for
chromosome aberrations. However, there was no effect under the utilized test conditions when
gaps were excluded from consideration. There was no evidence of structural chromosome
aberration induction by dibromochloromethane in the absence of exogenous metabolic activation.
Dibromochloromethane was tested in the mouse lymphoma assay as part of an
international collaborative program under the auspices of the Japanese Ministry of Health and
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Welfare (Sofuni et al., 1996). Dibromochloromethane yielded clearly positive results with or
without exogenous metabolic activation in two laboratories.
DeMarini et al. (1997) investigated the role of glutathione S-transferase activity in the
mutagenicity of dibromochloromethane in Salmonella typhimurium. Strains of Salmonella
utilized in this investigation included RSJ100, which is engineered to express the rat glutathione
S-transferase theta 1-1 (GSTT1-1) gene and TPT100, which has the GSTT1-1 gene inserted in a
non-functioning orientation. Dibromochloromethane (400 ppm) induced an 85-fold increase in
revertant colonies in the RSJ100 strain of Salmonella compared to background revertant
formation. The mutational spectra for dibromochloromethane-induced mutations at the hisG46
allele in strain RSJ100 were analyzed using the colony probe hybridization method. This analysis
revealed that 100% of the mutations were GC—>AT. A non-brominated dihalomethane,
dichloromethane, was tested in TA100 (which does not express GSTT1-1) for comparison. In
contrast to dibromochloromethane-induced mutations in RSJ100, only 15% of the mutations
induced by dichloromethane in TA100 were GC—>AT type mutations. This result suggests that
over-expression of GSTT1-1 in strain RSJ100 mediated the mutagenicity of
dibromochloromethane and induced a specific type of mutational lesion in Salmonella. Proposed
pathways of bioactivation of dibromochloromethane and other brominated trihalomethanes are
shown in Figure 4-2.
Landi et al. (1999) investigated the role of GSST1-1 in the mutagenicity of
dibromochloromethane in Salmonella by using one strain that expressed rat GSST1-1 (RSJ100)
and one strain that did not (TPT100). Mutagenicity of dibromochloromethane was assessed by
revertant colony formation with or without S9 metabolic activation. The addition of 800 ppm
dibromochloromethane greatly increased revertant numbers in the RSJ100 but not the TPT100
strain of Salmonella. Addition of the rat liver S9 fraction had no effect on the number of
revertants induced by dibromochloromethane exposure in either strain. These data provide
further support for the hypothesis that GSST1-1 plays a role in the mutagenicity of
dibromochloromethane. Additional experiments were conducted to investigate the effects of
exogenously added GSST1-1 on the mutagenic potency of dibromochloromethane. Red blood
cells (RBC), which express GSST1-1, were added to the experimental system to address this
question. RBC had no effect on results obtained with the TPT100 strain, but completely
suppressed the mutagenicity of dibromochloromethane in the RSJ100 strain. However, the
'protective' effect of RBC did not appear to be related to GSST1-1 activity, as this suppression
occurred even with the addition of RBC from individuals who do not express GSST1-1. The
underlying mechanism of RBC suppression of dibromochloromethane mutagenicity was not
investigated. The authors of this study speculated that tissues potentially exposed to
dibromochloromethane via the blood may be at less genotoxic risk (due to protection afforded by
the RBC) than tissues which are directly exposed to oral bromodichloromethane (such as tissues
in the gastrointestinal tract).
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In Vivo Assays
Fujie et al. (1990) analyzed chromosome aberrations in bone marrow from Long-Evans
rats (3/sex/dose) following oral (males only) or intraperitoneal (males and females) exposure to
dibromochloromethane. Oral administration was by gavage in saline for five consecutive days,
and the animals were sacrificed 18 hours after the last dose. Dibromochloromethane induced
dose-related increases in chromosome breaks. A more pronounced increase in clastogenic activity
was observed following a single intraperitoneal dose, with significant (p < 0.05) effects at 20.8
mg/kg. Regardless of the route, the predominant types of induced aberrations were chromatid
and chromosome breaks.
Hayashi et al. (1988) measured induction of micronucleated polychromatic erythrocytes in
ddY mice by intraperitoneal administration of dibromochloromethane at single doses of up to 500
mg/kg in corn oil. No evidence of clastogenicity was observed. However, the sampling time
utilized in this experiment was insufficient (U.S. EPA, 1994b). There was no clear evidence of
toxicity or cytotoxicity in the target tissue.
Ishidate et al. (1982) investigated the in vivo clastogenicity of dibromochloromethane in
ddY and MS mice and Wistar rats. Doses of 125 to 500 mg/kg-day were administered in olive oil
by intraperitoneal injection, and the animals were sacrificed at 18, 24, 30, 48, and 72 hours after
dosing. No significant induction of micronucleus formation was observed in either mice or rats.
Morimoto and Koizumi (1983) investigated the potential of dibromochloromethane to
induce sister chromatid exchanges in male ICR/SJ mice. Animals were given doses of 0, 25, 50,
100, or 200 mg/kg-day for four days by olive oil gavage. Dibromochloromethane produced a
roughly linear dose-dependent increase in sister chromatid exchange frequency. The increase was
statistically significant (p < 0.05) at 25 mg/kg-day. The authors noted that the concentrations
required to produce an increased incidence of sister chromatid exchange were on the order of
1,000 to 10,000 times higher than the concentrations typically found in drinking water.
Potter et al. (1996) dosed male F344 rats (4/dose) with 0.75 or 1.5 mmol/kg of
dibromochloromethane in 4% Emulphor® by gavage for 1, 3, or 7 days and investigated several
endpoints potentially related to kidney tumorigenesis. These doses corresponded to 156 or 312
mg/kg-day. No effect was observed when DNA strand breaks in the kidney were analyzed using
the alkaline unwinding procedure one day following treatment with a single dose of
dibromochloromethane. Because kidney tumors induced by some chemicals in male rats have
been related to the formation of a2u-globulin rich hyaline droplets, kidney hyaline droplets were
also evaluated in all of the dosed rats. Binding to a2u-globulin was not measured. No exposure-
related increase in hyaline droplets was found. Changes in kidney tubule cell proliferation were
assessed by in vivo incorporation [3H]-thymidine. No statistically significant effect of
dibromochloromethane exposure on this endpoint following was noted exposures of up to 7 days
duration.
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Stocker et al. (1997) investigated the in vivo genotoxicity of dibromochloromethane in an
unscheduled DNA synthesis assay in the livers of dibromochloromethane treated rats. Male
Sprague-Dawley rats (4 animals per group) were administered a single dose of 0 (control), 600 or
2000 mg/kg via gavage in aqueous 1% methylcellulose. These doses were selected by the authors
to correspond to 30% and 100% of the calculated maximum tolerated dose (MTD) for this
compound. Analysis of hepatocytes for unscheduled DNA synthesis was conducted 2 and 14
hours after treatment. There was no evidence of increased DNA synthesis in hepatocytes from
rats treated with any tested dose of dibromochloromethane.
Sekihashi et al. (2002) obtained positive results for dibromochloromethane genotoxicity
using the Comet assay. The authors indicated that doses were selected to avoid confounding of
the results by cytotoxicity. In Wistar rats, positive (statistically significant differences in mean
migration) results were obtained for stomach, colon, liver, kidney, bladder, or lung tissues
removed 8 or 24 hours following administration of 200 mg/kg oral dose of. using. In ddY mice,
positive results were obtained for liver and brain samples harvested 8 or 24 hours, respectively,
after administration of a 400 mg/kg oral dose. Although a statistically significant increase in
migration was also noted for the eight hour colon sample, the study authors did not identify this
finding as a positive response.
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Table V-ll Summary of Mutagenicity, Genotoxicity, and Neoplastic Transformation Data
for Dibromochloromethane
Endpoint
Assay System
Results
(with/without
activation)1*
References
In Vitro Studies
Gene mutation
Salmonella typhimurium
TA1003
NT/+
Simmon and Tardiff (1978)
TA100b
-/+
Ishidate et al. (1982)
TA98, TA100, TA1535,
TA1537b
-/-
NTP (1985)
TA1535, TA1537b
TA98, TA100b
+
+ +
Varma et al. (1988)
RSJ100
NT/+
DeMarini et al. (1997)
RSJ100
TPT100
+/+
-/-
Landi et al. (1999)
Mouse lymphoma cells3
NT/+
McGregor et al. (1991)
Mouse lymphoma cells
+/+
Sofuni et al. (1996)
Chromosome aberration
Chinese hamster fibroblastsb
+/-
Ishidate et al. (1982)
Chinese hamster ovary cellsb
-/-
Loveday et al. (1990)
Chinese hamster lung
fibroblasts3
-/+
(see text)
Matsuoka et al. (1996)
DNA damage
Saccharomyces cerevisiae3
-/+
Nestmann and Lee (1985)
SOS chromotest
S. typhimurium fluctuation
test
+/+
LeCurieux et al. (1995)
Sister chromatid
exchange
Human lymphocytes3
NT/+
Morimoto and Koizumi (1983)
Human lymphocytes3
+/NT
Sobti (1984)
Rat liver cellsb
+/NT
Sobti (1984)
Chinese hamster ovary cellsb
+/-
Loveday et al. (1990)
Rat erythroblastic leukemia
cells
-7+
Fujie et al. (1993)
In Vivo Studies
Micronuclei
Mouse bone marrow cells
-
Ishidate et al. (1982)
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Table V-ll (cont.)
Endpoint
Assay System
Results
(with/without
activation)1*
References
Mouse bone marrow cells
-
Hayashi et al. (1988)
Chromosome
aberrations
Rat bone marrow cells
+
Fujie et al. (1990)
Rat bone marrow cells
+
Fujie et al. (1990)
Sister chromatid
exchange
Mouse bone marrow cells
+
Morimoto and Koizumi (1983)
DNA damage
Rat kidney cells
-
Potter et al. (1996)
Rat stomach, colon, liver,
kidney, bladder, lung tissue
+
Sekihashi et al. (2002)
Mouse liver and brain tissue
+
Sekihashi et al. (2002)
Unscheduled DNA
synthesis
Rat hepatocytes
-
Stacker et al. (1997)
NT = Not Tested
a Assay was conducted in a closed system.
b Authors did not specify whether or not the assay was conducted in a closed system.
c Equivocal results reported.
d With/without activation applies to in vitro data only.
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3. Bromoform
The results of in vivo and in vitro tests conducted to evaluate the mutagenicity,
genotoxicity, and neoplastic transformation potential of bromoform are summarized in Table V-
12 at the end of this section.
In Vitro Assays
Simmon and Tardiff (1978) reported that nonactivated bromoform was mutagenic in S.
typhimurium strain TA100 when assayed as vapor in a desiccator. The minimum amount of
bromoform required to elicit a mutagenic response following addition to the desiccator was 570
|imol.
Ishidate et al. (1982) assayed the mutagenicity of bromoform in S. typhimurium strain
TA100 in the presence and absence of rat liver S9 fraction. Increased mutation frequencies were
observed only in the absence of S9 activation. In contrast, chromosomal aberrations in Chinese
hamster fibroblasts were observed in the presence, but not the absence, of S9 fraction. The
concentrations tested in these assays were not reported.
Maddock and Kelly (1980) reported that bromoform did not induce an increase in sister
chromatid exchanges when toadfish leukocytes were exposed to concentrations of 0.4 to 400 |iM.
Herren-Freund and Pereira (1986) assessed the initiating activity of bromoform using the
rat liver GGT-foci assay. The authors reported that a 250 mg/kg (1 mmol/kg) oral dose in an
unspecified vehicle did not initiate GGT-foci in this test.
NTP (1989a) evaluated the genotoxic potential of bromoform in multiple test systems.
Concentrations of 0.04 to 13 |imol/plate (10 to 3,333 |ig/plate) produced no evidence of
mutagenicity in S. typhimurium strains TA1535 or TA1537, when assayed with or without
exogenous metabolic activation by rat or hamster liver S9 fraction. Equivocal evidence of
mutagenicity was noted in strain TA100 without activation, and in strains TA97 and TA98 in the
presence of liver microsomes prepared from Aroclor-induced Syrian hamsters. Exposure of
mouse L5178Y cells to bromoform concentrations greater than or equal to 2,300 |iM in the
absence of S9 activation or S9-activated concentrations of at least 300 |iM with S9 activation
resulted in forward mutations at the thymidine kinase (tk) locus. One of two laboratories
conducting the assays reported increased sister chromatid exchanges (SCE) in CHO cells exposed
to 3,800 |iM bromoform in the absence of exogenous activation. Neither laboratory observed
increased incidence of SCE in the presence of S9. S9-activated bromoform did not induce
chromosome aberrations in CHO cells; results for SCE and chromosome aberrations in the
absence of exogenous activation were equivocal.
Zeiger (1990) found that bromoform was mutagenic in S. typhimurium strain TA98 when
tested as a vapor in a closed system, but not when tested in an open system using a preincubation
protocol. Positive results were observed at levels of at least 114 |imol/desiccator, in the presence
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and absence of S9 prepared from rat or hamster liver. Bromoform was negative in the closed
system with strains TA100 and TA1538 with or without rat or hamster liver S9 fraction
Roldan-Arjona and Pueyo (1993) evaluated bromoform in the S. typhimurium Ara
forward mutation assay at concentrations up to 25 |imol/plate (6.3 mg/plate). A preincubation
protocol was employed for the assay. Although a clear dose-related response was observed in the
absence of activation, the results were classified as questionable because a doubling of
background levels was not achieved. There was no evidence of mutagenicity in the presence of
exogenous metabolic activation. Although no attempt was made to minimize volatilization of the
test compound, cytotoxicity at the high exposure level indicated that the test material reached the
cells.
DeMarini et al. (1997) investigated the role of glutathione S-transferase activity in the
mutagenicity of bromoform in Salmonella typhimurium. Strains of Salmonella utilized for this
investigation included RSJ100, which has been engineered to express the rat glutathione S-
transferase theta 1-1 (GSTT1-1) gene and TPT100, which has the GSTT1-1 gene inserted in a
non-functioning orientation. Exposure to 1,600 ppm bromoform induced a 95-fold increase in
revertant colonies in the RSJ100 strain of Salmonella compared to background revertant
formation. The mutational spectra for bromoform-induced mutations at the hisG46 allele in strain
RSJ100 were analyzed using the colony probe hybridization method. This analysis revealed that
96% of the mutations were GC—>AT transitions. Bromoform also induced a smaller percentage
(2.8%) of GC—>TA mutations. A non-brominated halomethane, dichloromethane, was used in S.
typhimurium strain TA100 (which does not express GSST1-1) for comparison. In contrast to
bromoform-induced mutations in RSJ100, only 15% of the mutations induced by dichloromethane
in TA100 were GC—>AT type mutations. This result suggests that over-expression of GSTT1-1
in strain RSJ100 mediated the mutagenicity of bromoform and induced a specific type of
mutational lesion in Salmonella. Proposed pathways for the bioactivation of bromoform and
other brominated trihalomethanes are illustrated in Figure V-2.
Landi et al. (1999) investigated the mutagenicity of bromoform in in vitro exposed
human lymphocytes from both glutathione-S-transferase theta positive (GSST1-1+) and negative
(GSST1-1-) individuals. Whole blood cultures were exposed to bromoform (10"2 to 10"4 M) and
assayed for DNA breaks with the COMET assay. The DNA-damaging potency of bromoform
was not significantly different in lymphocytes (the target cell for the COMET assay) from GSST1-
1+ and GSST1-1- individuals. However, lymphocytes do not express GSST1-1, even in GSST1-
1+ individuals, so interpretation of this data is problematic. When data were combined from both
genotypic groups, there was a weak but statistically significant induction of comets observed
following treatment with bromoform.
In Vivo Assays
NTP (1989a) studied the genotoxic potential of bromoform in several test systems. Adult
male Drosophila fed with a 1,000-ppm solution of bromoform exhibited increased frequency of
sex-linked recessive lethal mutations, but no significant effect on reciprocal translocations was
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observed. Intraperitoneal injection of mice with 200 to 800 mg/kg bromoform caused an increase
in sister chromatid exchange but not in chromosomal aberrations in bone marrow cells.
Fujie et al. (1990) analyzed chromosome aberrations in bone marrow from Long-Evans
rats (3/sex/dose) following oral (males only) or intraperitoneal (males and females) exposure to
bromoform. Oral administration was by gavage in saline for five consecutive days, and the animals
were sacrificed 18 hours after the last dose. Bromoform induced a dose-related increase in the
incidence of aberrant cells, with a significant (p < 0.01) increase at 253 mg/kg-day. A more
pronounced increase in clastogenic activity was observed following a single intraperitoneal dose,
with a significant (p < 0.05) effect at 25.3 mg/kg. Regardless of the route, the predominant types
of induced aberrations were chromatid and chromosome breaks.
Morimoto and Koizumi (1983) investigated the ability of bromoform and other
brominated trihalomethanes to induce sister chromatid exchanges in human lymphocytes in vitro
in the absence of S9 activation. All three brominated trihalomethanes caused a dose-dependent
increase in sister chromatid exchanges. Bromoform was more potent than bromodichloromethane
or dibromochloromethane. The increases were significant (p < 0.05) at concentrations greater
than or equal to 400 |iM, 400 |iM, and 80 |iM for bromodichloromethane, dibromochloro-
methane, and bromoform, respectively.
Potter et al. (1996) evaluated the effect of bromoform on incidence of DNA strand breaks
in the kidney. Male F344 rats received 0.75 or 1.5 mmol/kg of bromoform in 4% Emulphor® by
gavage for 1, 3, or 7 days. These doses corresponded to 190 or 379 mg/kg. No effect was
observed on strand breaks when evaluated using the alkaline unwinding procedure one day after a
single dose.
Stocker et al. (1997) investigated the in vivo genotoxicity of bromoform in the mouse
bone marrow micronuclei assay and by analysis of unscheduled DNA synthesis in the liver of
bromoform-treated rats. In the first assay, Swiss CD mice (5/sex/dose) were treated by gavage
with doses of 0, 250, 500, or 1,000 mg/kg bromoform dissolved in aqueous 1% methylcellulose.
Micronuclei analysis was conducted 24 and 48 hours after dosing, and was negative in all dose
groups. In the second assay, male Sprague-Dawley rats (4 animals/dose) received single doses of
0, 324 or 1,080 mg/kg bromoform by gavage in aqueous 1% methylcellulose. These doses were
selected by the authors to correspond to 30% and 100% of the calculated MTD for this
compound. Analysis of hepatocytes for unscheduled DNA synthesis was conducted 2 and 14
hours after treatment. There was no evidence of increased DNA synthesis in hepatocytes from
rats treated with any tested dose of bromoform.
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Table V-12 Summary of Mutagenicity, Genotoxicity, and Neoplastic Transformation Data
for Bromoform
Endpoint
Assay System
Results
(with/without
activation)1*
References
In Vitro Studies
Gene mutation
Salmonella typhimurium
TA100\ TA1535
NT/+
Simmon and Tardiff (1978)
TA1535, TA1537b
TA100
TA97, TA98
-/-
-/±c
±7-
NTP (1989a)
TA100b
-/+
Ishidate et al. (1982)
TA98
TA100, TA15383
+/+
-/-
Zeiger (1990)
S. typhimurium Ara
-/+c
Roldan-Arjona and Pueyo
(1993)
RSJ100
NT/+
DeMarini et al. (1997)
Mouse lymphoma cellsb
+/+
NTP (1989a)
Chromosome aberration
Chinese hamster fibroblastsb
+/-
Ishidate et al. (1982)
Chinese hamster ovary cellsb
-/±
NTP (1989a)
DNA damage
Human lymphocytes
NT/+
Landi et al. (1999)
Sister chromatid
Toadfish leukocytes3
NT/-
Maddock and Kelly (1980)
exchange
Human lymphocytesb
NT/+
Morimoto and Koizumi (1983)
[Chinese hamster ovary cells
-/±
NTP (1989a)
Initiation
Rat liver GGT-foci assay
-
Herren-Freund and Pereira
(1986)
In Vivo Studies
Micronuclei
Mouse bone marrow cells
-
Ishidate et al. (1982)
Mouse bone marrow cells
-
Hayashi et al. (1988)
Mouse bone marrow cells
-
Stacker et al. (1997)
Chromosome
aberrations
Mouse bone marrow cells
-
NTP (1989a)
Rat bone marrow cells (oral)
+
Fujie et al. (1990)
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Table V-12 (cont.)
Endpoint
Assay System
Results
(with/without
activation)1*
References
DNA damage
Rat renal cells
-
Potter et al. (1996)
Unscheduled DNA
synthesis
Rat hepatocytes
-
Stacker et al. (1997)
Sister chromatid
exchange
Mouse bone marrow cells
+
Morimoto and Koizumi (1983)
Mouse bone marrow cells
(ip)
+
NTP (1989a)
Sex-linked recessive
lethal mutations
Drosophila
+
NTP (1989a)
NT = Not Tested
a Assay was conducted in a closed system.
b Authors did not specify whether or not the assay was conducted in a closed system.
c Equivocal results obtained.
d With/without activation applies to in vitro assays only.
G. Carcinogenicity
1. Bromodichloromethane
NTP (1987) evaluated the and carcinogenic potential of bromodichloromethane in F344/N
rats in a two-year study oral exposure study. Additional details of this study are provided in
Section V.D. 1. Groups of male and female rats (50/sex/group) were administered
bromodichloromethane in corn oil via gavage at doses of 0, 50, or 100 mg/kg-day for 5
days/week for 102 weeks. All animals were examined grossly and microscopically for neoplastic
lesions. Survival of all dosed animals was comparable to or greater than the corresponding
control group. Mean body weights of high-dose make and female rats was decreased during the
last 1.5 years of the study. Body weight gains of high dose male and female rats were 86% and
70% of the corresponding vehicle control group, respectively. Statistically significant increases in
the incidences of neoplasms of the large intestine and kidney were observed in male and female
rats (Table V-13). The study authors noted that neoplasms of the large intestine and kidney are
uncommon tumors in F344/N rats based on historical control data for NTP studies. They
concluded that under the conditions of these 2-year gavage studies, clear evidence of carcinogenic
activity existed in male and female rats.
NTP (1987) also evaluated the potential toxic and carcinogenic effects of
bromodichloromethane in rats and mice in a two-year study oral exposure study. Additional
details of this study are provided in Section V.D. 1. Groups of male and female B6C3F, mice
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Table V-12 (cont.)
(50/sex/dose) were administered doses of 0, 25, or 50 mg/kg-day (males) or 0, 75, or 150 mg/kg-
day (females) for 5 days/week for 102 weeks. All animals were examined grossly and
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Table V-13 Tumor Frequencies in F344/N Rats and B6C3Ft Mice Exposed to
Bromodichloromethane in Corn Oil for 2 Years - Adapted from NTP (1987)
Animal
Tissue/Tumor
Tumor Frequency
Control
50 mg/kg
100 mg/kg
Male Rat
Large intestine
Adenomatous polyp
0/50
3/49
33/50b
Adenocarcinoma
0/50
11/4 9b
38/50b
Combined
0/50
13/49b
45/50b
Kidneya
Tubular cell adenoma
0/50
1/49
3/50
Tubular cell adenocarcinoma
0/50
0/49
10/50b
Combined
0/50
1/49
13/50b
Control
50 mg/kg
100 mg/kg
Female Rat
Large intestine
Adenomatous polyp
0/46
0/50
7/47b
Adenocarcinoma
0/46
0/50
6/47b
Combined
0/46
0/50
12/47b
Kidney
Tubular cell adenoma
0/50
1/50
6/5 0b
Tubular cell adenocarcinoma
0/50
0/50
9/5 0b
Combined
0/50
1/50
15/50b
Control
25 mg/kg
50 mg/kg
Male Mouse
Kidneyd
Tubular cell adenoma
1/46
2/49
6/50
Tubular cell adenocarcinoma
0/46
0/49
4/50
Combined
1/46
2/49
9/5 0b
Control
75 mg/kg
150 mg/kg
Female Mouse
Liver
Hepatocellular adenoma
1/50
13/48b
23/50b
Hepatocellular carcinoma
2/50
5/48
10/50b
Combined
3/50
18/48b
29/50b
a One rat died at week 33 in the low-dose group and was eliminated from the cancer risk calculation.
b Statistically significant at p<0.05, compared to controls.
c Intestine not examined in four rats from control group and three rats from high-dose group.
d In the control group, two mice died during the first week, one mouse died during week, nine and one escaped in
week 79. One mouse in the low-dose group died in the first week. All of these mice were eliminated from the
cancer risk calculations.
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microscopically for neoplastic lesions. Survival of dosed male mice was comparable to the
corresponding control group. Survival of dosed and vehicle control females was decreased after
week 84 as a result of ovarian abscesses. Body weight gain in high-dose males was decreased by
13% when compared to the vehicle control group. Body weight gain in low- and high-dose
females was reduced by 25% and 55%, respectively. Statistically significant increases were
observed in the incidences of neoplasms of the kidney in male mice and the liver in female mice
(Table V-13). The study authors noted that neoplasms of the kidney are uncommon in B6C3F,
mice based on NTP historical control data. They concluded that under the conditions of these 2-
year gavage studies, clear evidence of carcinogenic activity existed in male and female mice.
Tumasonis et al. (1987) exposed groups of 58 male and female Wistar rats to
bromodichloromethane in drinking water from weaning until death occurred in all of the animals
(approximately 185 weeks). The exposure level was 2,400 mg/L for 72 weeks and was reduced
to 1,200 mg/L for the remaining 113 weeks. Based on a graph presented by the authors, the
average dose over the course of the experiment was probably about 150 mg/kg-day for females
and about 100 mg/kg-day for males. Exposed animals of both sexes gained significantly less
weight (approximately 30 to 40%) than control animals. There was a statistically significant (p <
0.01) increase in the incidence of hepatic neoplastic nodules in exposed females compared to
control females (32% versus 0%). Significant increases were also reported for the occurrence of
hepatic adenofibrosis (12% versus 0%) and lymphosarcoma (17% versus 11%) in females. No
statistically significant increase in the incidence of any tumor was reported in males. Two males
and one female among the treated animals were observed to have renal adenoma or carcinoma,
while no renal tumors were observed in the controls. Statistically significant decreases in the
incidence of mammary tumors and pituitary tumors in females and lymphosarcomas in males were
observed.
Aida et al. (1992b) administered bromodichloromethane to Slc:Wistar rats
(40/sex/treatment group and 70/sex/controls) at dietary levels of 0%, 0.014%, 0.055%, or 0.22%
for up to 24 months. The test material was microencapsulated and mixed with powdered feed.
Based on the mean food intakes, the mean doses were 0, 6.1, 25.5, or 138.0 mg/kg-day for males
and 0, 8.0, 31.7, or 168.4 mg/kg-day for females. The only neoplastic lesions observed were
three cholangiocarcinomas and two hepatocellular adenomas in the high-dose females, one
hepatocellular adenoma in a control female, one cholangiocarcinoma in a high-dose male, and one
hepatocellular adenoma each in a low-dose male and a high-dose male. Based on these results,
the study authors concluded that there was no clear evidence that microencapsulated
bromodichloromethane administered in the diet was carcinogenic in Wistar rats.
Voronin et al. (1987) assessed the carcinogenic potential of bromodichloromethane in
male and female CBA x C57B1/6 mice. Groups of mice (50-55/sex/concentration) were exposed
to bromodichloromethane provided in drinking water at concentrations of 0.04, 4.0, or 400 mg/L
for 104 weeks. Untreated control groups of 75 male and 50 female mice were also included in the
study design. No significant differences were observed in total tumor incidence when evaluated
by Chi square analysis. The study authors concluded that, under the conditions of this bioassay,
bromodichloromethane was not carcinogenic in mice.
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Theiss et al. (1977) examined the carcinogenic potential of bromodichloromethane in
Strain A mice (6 to 8 weeks old). Male animals (20 mice/group) were injected intraperitoneally
up to three times weekly over a period of 8 weeks. Three dose levels (20, 40, or 100 mg/kg
bromodichloromethane) were used with concurrent positive and negative control groups that
contained 20 animals each. Mice were sacrificed 24 weeks after the first injection, and the
frequency of lung tumors in each test group was compared with vehicle-treated controls. No
statistically significant increase in the incidence of lung tumors/mouse was reported.
Melnick et al. (1998) investigated the mechanistic relationship between liver toxicity and
tumorigenicity of bromodichloromethane. Female B6C3F, mice (10 animals per group) were
treated with bromodichloromethane in corn oil via gavage for 3 weeks (5 days/week). Doses of
bromodichloromethane used in this study were 0 (vehicle only), 75, 150, or 326 mg/kg-day. A
significant dose-related increase in absolute liver weight and liver weight/body weight ratio was
noted for the 150 and 326 mg/kg-day dose groups. Serum ALT activity was significantly
increased in the two highest dose groups and serum SDH activity was elevated at all doses tested.
At necropsy, there was clear evidence of hepatocyte hydropic degeneration in animals treated
with 150 and 326 mg/kg-day. BrdU was administered to the animals during the last 6 days of the
study, and hepatocyte labeling index (LI) analysis was conducted. The two highest (150 and 326
mg/kg-day) doses resulted in significantly elevated hepatocyte proliferation as measured by the LI.
Using the Hill equation model, these authors compared the dose response for liver toxicity
(enzyme and LI data) and tumorigenicity (data from previously published NTP sponsored
bioassays) for bromodichloromethane. This analysis indicated that the shape of the dose response
as well as the Hill exponents were different for liver toxicity and tumorigenicity. It was concluded
that the results of this comparison do not support a causal relationship between liver
toxicity/reparative hyperplasia and tumor development.
George et al. (2002) evaluated the carcinogenicity of bromodichloromethane in male
F344/N rats (78 animals/dose) exposed to the compound via drinking water for 104 weeks.
Nominal concentrations of 0.07, 0.35, or 0.70 g/L were administered in drinking water containing
0.25% Emulphor®. The vehicle control solution consisted of 0.25% Emulphor®. The study
authors indicated that testing of higher concentrations was prevented by refusal of the test animals
to drink solutions containing more than 0.7 g/L. Six animals per exposure concentration were
sacrificed at 13, 26, 52, and 78 weeks for gross observation and histopathological examination of
the thyroid, liver, stomach, duodenum, jejunum, ileum, colon, rectum, spleen, kidneys, urinary
bladder, and testes. A complete rodent necropsy was performed at terminal sacrifice and
representative samples of the tissues listed above were examined microscopically. A complete
pathological examination was performed on five rats from the high dose group. Serum profiles of
LDH, ALT, ALP, AST, SDH, BUN, total protein, creatine, and total antioxidant activities were
determined at 26, 52, and 104 weeks. Hepatocyte and renal tubular cell proliferation were
measured at each sacrifice by bromodeoxyuridine labeling.
The measured drinking water concentrations of bromodichloromethane were 0.06, 0.38,
and 0.76 g/L. When corrected for loss of bromodichloromethane as a result of volatility,
instability, or adsorption to glass surfaces during treatment, the corresponding administered
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concentrations were 0.06, 0.33, and 0.62 g/L. Based on measured water consumption, these
levels correspond to mean daily doses for the entire study of 3.9, 20.6, and 36.3 mg/kg-day as
calculated by the study authors. Mean daily doses of 6.4, 32.6, and 58.9 mg/kg-day were
calculated for the first 13 weeks of the study when the growth rate of the test animals was
highest. No significant differences were observed among groups for feed consumption or
survival. Twenty-one to 22 unscheduled deaths were observed in each treatment group.
Mononuclear cell leukemia was seen in all dose groups and was reported to be the primary cause
of morbidity and mortality prior to 104 weeks. Exposure to bromodichloromethane did not affect
the growth rate of test animals when compared to the control. Kidney weight was significantly
depressed at the high dose and a significant negative trend was observed for relative kidney
weight. No significant changes were observed in clinical chemistry parameters. Observed
nonneoplastic changes in the liver (e.g., biliary fibrosis, bile duct inflammation, and chronic
inflammation) were considered to be age-related background changes, since neither the incidence
nor severity of the lesions differed from the control values. Bromodichloromethane had no effect
on hepatocyte proliferation as measured by bromodeoxyuridine labeling. Renal tubular cell
hyperplasia was significantly decreased in the 3.9 mg/kg-day group and significantly increased in
the 36.3 mg/kg-day group (15.8%) relative to the control value (8.7%).
The absence of effect on body weight and other examined endpoints suggests that a
maximum toxic dose may not have been achieved in this study. However, the dosing regimen
used by George et al. (2002) was sufficient to increase the incidence of hepatocellular neoplasia
(Table V-14). The data for hepatic tumors indicate a biphasic pattern of dose-response. The
prevalence and multiplicity of hepatocellular adenoma and combined hepatocellular adenoma and
carcinoma were significantly increased at 3.9 mg/kg-day, nonsignificantly increased at 20.6
mg/kg-day, and comparable to the control values at 36.3 mg/kg-day. The prevalence and
multiplicity of hepatocellular carcinoma were increased at 20.6 mg/kg-day when compared to
control values, but the response did not reach statistical significance. The underlying basis for the
biphasic response is unknown, but the study authors noted that the observed pattern of response
could be explained by inhibition of the hepatic metabolism of bromodichloromethane by the
compound itself. Exposure to bromodichloromethane decreased the prevalence of basophilic
(control, 67%; 3.6 mg/kg-day, 62.2%; 20.6 mg/kg-day, 46%; 36.6 mg/kg-day, 34.7%) and clear
cell (17.8%), 2.2%), 2.1%, 4.1%>) altered foci of cells (AFC) in a dose-dependent manner, but had
no significant effect on the prevalence of eosinophilic AFCs when compared to the controls. The
decreases in prevalence were statistically significant at the mid and high doses for basophilic AFCs
and at all doses for clear cell AFCs. Exposure to bromodichloromethane had no significant effect
on the prevalence of renal tubular adenomas or carcinomas (Table V-14). One renal tubular
adenoma was observed in the 3.6 mg/kg-day group and two tumors were observe in the 36.3
mg/kg-day. The historical incidence of renal tubular adenomas in male F344/N rats is very low
(2/327 or 0.6%>), as determined from control groups in NTP drinking water studies (NTP, 2002;
no data were reported from the study laboratory). Therefore, the occurrence of these tumors in
the present study may be of biological significance. No increased incidences of neoplasia were
evident in the five high dose animals selected for a histopathological examination of all organs.
On the basis of the increased prevalence and multiplicity of hepatocellular neoplasms in the 3.9
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and 20.6 mg/kg-day groups, the study authors concluded that bromodichloromethane was
carcinogenic in male F344/N rats under the
Table V-14 Hepatic and Renal Tumors in Male F344/N Rats Administered
Bromodichloromethane in the Drinking Water for Two Years
Tumor Type
Mean Daily Dose of Bromodichloromethane (mg/kg-day)
Vehicle Control
3.9
20.6
36.3
Liver
Hepatocellular adenoma
1/45 (2.2%)a
0.02 ± 0.02b'c
7/45 (15.5%)*
0.16 ±0.04*
3/48 (6.2%)
0.06 ± 0.02
2/49 (4.1%)
0.04 ± 0.02
Hepatocellular carcinoma
1/45 (2.2%)
0.02 ± 0.02
1/45 (2.2%)
0.02 ±0.01
4/48 (8.3%)
0.10 ±0.03
2/49 (4.1%)
0.04 ± 0.02
Hepatocellular adenoma and
carcinoma (combined)
2/45 (4.4%)
0.04 ± 0.02
8/45 (17.8%)*
0.19 ±0.00*
7/48 (14.6%)
0.17 ±0.04
4/49 (8.2%)
0.08 ±0.28
Kidney
Tubular cell adenoma
0/46 (0%)
1/45 (2.2%)
0/51 (0%)
2/44 (4.5%)
Tubular cell carcinoma
0/46 (0%)
0/45 (0%)
0/51 (0%)
0/44 (0%)
Tubular cell adenoma or
carcinoma (combined)
0/46 (0%)
1/45 (2.2%)
0/51 (0%)
2/44 (4.5%)
Source: George et al. (2002)
* Statistically significant when compared to the control value, p<0.05
a Prevalence (percentage of animals with tumor)
b Multiplicity, number of tumors per animal
c Mean ± standard deviation
conditions of the bioassay. A source of uncertainty in this conclusion is lack of knowledge on the
biological mechanism underlying the biphasic dose-response observed for hepatic tumors.
George et al. (2002) also evaluated the carcinogenicity of bromodichloromethane in male
B6C3F, mice (78 animals/dose) exposed via drinking water for 100 weeks. Nominal
concentrations of 0.05, 0.25, or 0.50 g/L were administered in drinking water containing 0.25%
Emulphor®. The vehicle control solution consisted of 0.25% Emulphor®. Seven animals per
exposure concentration were sacrificed at 13, 26, 52, and 78 weeks for gross observation and
histopathological examination of the liver, stomach, duodenum, jejunum, ileum, colon, rectum,
spleen, kidneys, urinary bladder, and testes. A complete rodent necropsy was performed at
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terminal sacrifice and representative samples of the tissues listed above were examined
microscopically. A complete pathological examination was performed on five rats from the high
dose group. Serum profiles of LDH, ALT, ALP, AST, SDH, BUN, total protein, creatine, and
total antioxidant activities were determined at 26, 52, and 100 weeks. Hepatocyte and renal
tubular cell proliferation were measured by bromodeoxyuridine labeling at each sacrifice.
The measured drinking water concentrations of bromodichloromethane were 0.06, 0.30,
and 0.55 g/L. When corrected for loss of bromodichloromethane as a result of volatility,
instability, or adsorption to glass surfaces during treatment, the corresponding administered
concentrations were 0.06, 0.28, and 0.49 g/L. Based on measured water consumption, these
levels correspond to mean daily doses of 8.1, 27.2, and 43.4 mg/kg-day as calculated by the study
authors. Water consumption was significantly reduced at the mid- and high doses; the study
authors attributed the reduced intake to taste aversion. No significant differences were observed
among groups for feed consumption or survival. Exposure to bromodichloromethane did not
affect the growth rate of test animals when compared to the control. Kidney weight was
significantly depressed at 27.2 and 43.4 mg/kg-day when compared to the control values. No
significant changes were observed in clinical chemistry parameters. Mild, treatment-related
nonneoplastic hepatic lesions were observed in the 27.2 and 43.4 mg/kg-day dose groups (identity
and prevalence not reported). Increased incidences of hepatocellular karyomegaly and necrosis
with inflammation (prevalence and severity not reported) were not dose-related. The prevalence
of renal tubular hyperplasia was 3%, 0%, 6% and 0% for the vehicle control, 8.1, 27.2, and 43.4
mg/kg-day groups, respectively. Other observed preneoplastic and neoplastic lesions (identity and
prevalence not reported) were considered background events for the male B6C3F, mouse. BrdU
labeling index in hepatocytes and renal tubular cells was not altered at any time point.
Hepatocellular adenomas and carcinomas were observed in all treatment groups. Neither the
prevalence nor multiplicity of these tumors was significantly increased by exposure to
bromodichloromethane. Renal tubular cell neoplasia was not observed in any treatment group.
No increased incidences of neoplasia were evident in the five high dose animals subject to a full
histopathological examination. On the basis of these data, the study authors concluded that
bromodichloromethane was not carcinogenic to male mice under the conditions employed in this
study. However, in the absence of compound-related effects on body weight or other toxicologic
endpoints, it is not evident that an adequately high dose was tested in this study.
De Angelo et al., (2002) evaluated the ability of bromodichloromethane administered in
drinking water to induce aberrant crypt foci (ACF), putative early preneoplastic lesions, in the
colons of male F344/N rats. Groups of weanling rats (6 animals/group) were exposed to distilled
water, 0.25% Alkamuls EL-620®, or 0.7 g/L bromodichloromethane in 0.25% Alkamuls EL-620
for 13 weeks. A single intraperitoneal injection of 30 mg/kg azoxymethane (AOM) served as the
positive control. Body weight and water consumption were measured twice during the first week
of the study and once per week thereafter. Colons were collected at study termination, fixed,
stained with 0.2% methylene blue, divided into three equal segments, and scanned for ACF. The
measured concentration of bromodichloromethane averaged 0.64 ± 0.06 mg/L (mean and
standard error) over the course of the study. When adjusted for volatilization and adherence to
glass, the corrected concentration was 0.51 mg/L. Water consumption was significantly reduced
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(38%) in the bromodichloromethane exposure group when compared to the deionized water
control. The average daily dose was 45 mg/kg-day as calculated by the study authors. Average
terminal body weight of the rats exposed to bromodichloromethane was within 10% of the control
values. No ACF were observed in colons from control animals. ACF were observed in five of six
colons from bromodichloromethane exposed animals. The total number of ACF (30) and number
of aberrant crypts per focus (3.33 ± 0.47) were significantly increased relative to the combined
deionized water and vehicle controls. All observed ACF were located in the distal (rectal)
segment of the colon. For comparison, 807 ACF and 4.95 ± 0.25 crypts per focus (mean and
standard error) were observed in the AOM positive control group. Eight percent, 42% and 50%
of the ACF induced by AOM were located in the proximal, middle, and distal segment of the
colon, respectively. The study authors noted that ACF induced by bromodichloromethane do not
to progress to neoplasia, as judged by the absence of colon neoplasms in the two-year cancer
study conducted by George et al. (2002).
De Angelo et al. (2002) also evaluated the ability of bromodichloromethane administered
in drinking water to induce ACF in the colons of male B6C3F, (6 animals/group) and A/J mice (9
animals/group; sex not specified). Mice of the A/J strain are sensitive to chemical induction of
ACF. Test animals were exposed to distilled water, 0.25% Alkamuls EL-620®, or 0.5 g/L
bromodichloromethane in 0.25% Alkamuls EL-620 for 13 weeks (both strains) or 30 weeks (A/J
mice only). A single intraperitoneal injection of 50 mg/kg 4-aminobiphenyl or 10 mg/kg
azoxymethane (AOM) served as the positive controls for the B6C3F1 and A/J strains,
respectively. Body weight and water consumption were measured twice during the first week of
the study and once per week thereafter. Colons were collected at study termination, fixed, stained
with 0.2% methylene blue, divided into three equal segments, and scanned for ACF. The study
report did not provide results for measured concentration of bromodichloromethane in drinking
water solutions or an estimated dose. No differences were observed in between the control and
any treatment group for body weight or water and feed consumption. ACF development was not
observed in the colons of B6C3FX mice treated with bromodichloromethane in the drinking water
or injected with 4-aminobiphenyl. Bromodichloromethane did not induce ACF in A/J mice.
Injection of A/J mice with the positive control compound AOM induced 47.4 ± 4.9 ACF/cm2
(mean and standard error) and 7.2 ± 1.1 tumors/cm2 after 13 weeks and 17.8 ± 2.6 tumors/cm2
after 30 weeks of treatment. In comparison, 807 ACF and 4.95 ± 0.25 crypts per focus were
observed in the AOM positive control group.
2. Dibromochloromethane
NTP (1985) administered dibromochloromethane at doses of 0, 40, or 80 mg/kg-day (in
corn oil) to groups of 50 male and 50 female F344/N rats via gavage 5 times/week for 104 to 105
weeks. Survival of dosed male and female rats was comparable to that of the vehicle-control
groups. High-dose males had lower body weights when compared with the vehicle control.
Compound-related nonneoplastic lesions (fatty metamorphosis and ground-glass cytoplasmic
changes) were found in the livers of both sexes (See section V.D. 1). Nephrosis was observed in
female rats. No statistically significant increase in the incidence of any neoplastic lesion was
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observed. Based on the results of this study, the authors concluded that there was no evidence of
carcinogenicity in rats administered dibromochloromethane.
NTP (1985) administered dibromochloromethane (in corn oil) to groups of 50 male and
50 female B6C3F, mice via gavage 5 times/week for 104 to 105 weeks. The administered doses
were 0, 50, or 100 mg/kg-day. Survival of female mice was comparable to that of the vehicle-
control group. High-dose male mice, however, had lower survival rates than the vehicle controls.
At week 82, nine high-dose male mice died of an unknown cause. An inadvertent overdose of
dibromochloromethane given to low-dose male and female mice at week 58 killed 35 male mice,
but apparently did not affect the females. The low-dose male mouse group was, therefore,
considered to be unsuitable for analysis of neoplasms. Compound-related nonneoplastic lesions
were found primarily in the livers of male mice (hepatocytomegaly, necrosis, fatty metamorphosis)
and female mice (calcification and fatty metamorphosis). Nephrosis was observed in male mice.
In females, a statistically significant increase in the incidence of hepatocellular adenomas and
adenomas and carcinomas combined was observed in the high-dose group. In male mice, a
statistically significant increase in the incidence of hepatocellular carcinomas and adenomas and
carcinomas combined was observed in the high-dose group. A summary of the incidence of these
tumors is presented in Table V-15. A negative trend in the incidence of malignant lymphomas
was evident in dibromochloromethane-exposed male mice when compared to the vehicle control.
The study authors concluded that the results of this study provided equivocal evidence of
dibromochloromethane carcinogenicity in male B6C3F, mice and some evidence of
carcinogenicity in female B6C3F, mice.
In other bioassays, Voronin et al. (1987) observed no significant tumor increases in
CBAxC57Bl/6 mice (50/sex/dose) treated with dibromochloromethane in the drinking water at
concentrations of 0, 0.04, 4.0, or 400 mg/L (approximately 0, 0.008, 0.76, or 76 mg/kg-day) for
104 weeks. In an unpublished report of a two-year dietary study, Tobe et al. (1982) reported no
increase in gross tumors in male rats dosed with up to 210 mg/kg-day or female rats treated with
up to 350 mg/kg-day.
Melnick et al. (1998) exposed female B6C3F, mice (10/dose) to dibromochloromethane in
corn oil via gavage for 3 weeks (5 days/week). The doses of dibromochloromethane in this study
were 0 (vehicle only), 50, 100, 192, or 417 mg/kg-day. The corresponding time-weighted doses
are 0, 37, 71, 137, and 298 mg/kg-day. No treatment-related signs of overt toxicity were
observed during the study. Body weight and water intake were not significantly altered at any
dose tested. However, a statistically significant and dose-related increase in liver weight/body
weight ratio was seen in the 100, 192 and 417 mg/kg-day dose groups. Serum ALT activity was
significantly increased in the two highest dose groups. The activity of serum SDH was
significantly elevated at all doses tested except 50 mg/kg-day. However, the increase in activity
(shown graphically) was very small relative to the control at the 100 and 192 mg/kg-day doses.
At necropsy, there was clear evidence of hepatocyte hydropic degeneration in the 192 and 417
mg/kg-day dose groups. BrdU was administered to the animals during the last 6 days of the
study, and hepatocyte labeling index (LI) analysis was conducted. Only the highest dose tested
(417 mg/kg-day) resulted in significantly elevated hepatocyte proliferation as measured by the LI.
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Using the Hill equation model, these authors compared the dose response for liver toxicity
(enzyme and LI data) and tumorigenicity (data from previously conducted NTP bioassays) for
dibromochloromethane. This analysis indicated that the shape of the dose response as well as the
Hill exponents were different for liver toxicity and tumorigenicity. It was concluded that the
results of this comparison do not support a causal relationship between liver toxicity/reparative
hyperplasia and tumor development.
Table V-15 Frequencies of Liver Tumors in B6C3Ft Mice Administered
Dibromochloromethane in Corn Oil for 105 Weeks - Adapted from NTP (1985)
Treatment
Sex
Adenoma
Carcinoma
Adenoma or Carcinoma
(mg/kg-day)
(combined)
Vehicle Control
M
14/50
10/50
23/50
F
2/50
4/50
6/50
50
M
a
F
4/49
6/49
10/49
100
M
10/50
19/50b
27/50c
F
1 l/50b
8/50
19/50d
a Male low-dose group was inadequate for statistical analysis.
b p < 0.05 relative to controls.
c p < 0.01 (life table analysis); p = 0.065 (incidental tumor test) relative to controls.
d p < 0.01 relative to controls.
De Angelo et al. (2002) evaluated the ability of dibromochloromethane administered in
drinking water to induce aberrant crypt foci (ACF), putative early preneoplastic lesions, in the
colons of male F344/N rats. Groups of weanling rats (6 animals/group) were exposed to distilled
water, 0.25% Alkamuls EL-620®, or 0.9 g/L dibromochloromethane in 0.25% Alkamuls EL-620
for 13 weeks. A single intraperitoneal injection of 30 mg/kg azoxymethane (AOM) served as the
positive control. Body weight and water consumption were measured twice during the first week
of the study and once per week thereafter. Colons were collected at study termination, fixed,
stained with 0.2% methylene blue, divided into three equal segments, and examined for ACF. The
measured concentration of dibromochloromethane averaged 0.80 ± 0.05 mg/L (mean and
standard error) over the course of the study. When adjusted for volatilization and adherence to
glass, the corrected concentration was 0.63 mg/L. Water consumption was significantly reduced
(32%) in the dibromochloromethane exposure group when compared to the deionized water
control. The average daily dose of dibromochloromethane was 60 mg/kg-day as calculated by the
study authors. Average terminal body weight of the rats exposed to dibromochloromethane was
within 10%) of the control values. No ACF were observed in colons from control animals. ACF
were observed in three of six colons from dibromochloromethane-exposed animals. The total
number of ACF (17) and number of aberrant crypts per focus (2.43 ± 0.61) were significantly
increased relative to the combined deionized water and vehicle controls. Fourteen percent of the
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observed ACF were located in the middle segment of the colon and 86% were located in the distal
(rectal) segment. In comparison, 807 ACF and 4.95 ± 0.25 crypts per focus were observed in the
AOM positive control group. Eight percent, 42% and 50% of the ACF induced by AOM were
located in the proximal, middle, and distal segment of the colon, respectively.
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3. Bromoform
Groups of 50 male B6C3F, mice were exposed to bromoform via gavage (corn oil) at
doses of 0, 50, or 100 mg/kg-day of bromoform for 103 weeks (5 days/week) (NTP, 1989a).
Groups of 50 female mice received doses of 0, 100, or 200 mg/kg-day bromoform by the same
protocol. At termination, all animals underwent gross necropsy and thorough histological
examinations of tissues. Survival in both treated female groups was reduced; however, the
authors attributed this reduction in survival partly to utero-ovarian infection. A statistically
significant increase in the incidence of thyroid follicular cell hyperplasia was noted in high-dose
females; however, there were no statistically significant increases in the incidence of any
neoplastic lesion in any dose group compared to controls. Based on the results of this study, the
NTP (1989a) concluded there was no evidence of carcinogenic activity of bromoform in male or
female mice.
In a similar experiment, groups of 50 male and 50 female F344/N rats were administered
bromoform via gavage in oil at doses of 0, 100, or 200 mg/kg-day for 5 days/week for 103 weeks
(NTP 1989a). At termination, all animals were necropsied, and a thorough histological
examination of tissues was performed. Adenomatous polyps or adenocarcinomas of the large
intestine were noted in three high-dose male rats, eight high-dose female rats, and one low-dose
female rat (Table V-16). Although the number of tumors found was small, the incidence was
considered to be significant because these intestinal tumors are very rare in the rat. The NTP
concluded that there was some evidence for carcinogenic activity in male rats and clear evidence
in female rats.
Table V-16 Tumor Frequencies in the Large Intestine of F344/N Rats Exposed to
Bromoform in Corn Oil for 2 Years - Adapted from NTP (1989a)
Tumor
Tumor Frequency
Male rat
Control
100 mg/kg
200 mg/kg
Adenocarcinoma
0/50
0/50
1/50
Polyp (adenomatous)
0/50
0/50
2/50
Female rat
Control
100 mg/kg
200 mg/kg
Adenocarcinoma
0/48
0/50
2/50
Polyp (adenomatous)
0/48
1/50
6/50
Theiss et al. (1977) examined the carcinogenic activity of bromoform in Strain A mice.
Male animals (6 to 8 weeks old, 20 mice/group) were injected intraperitoneally up to three times
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weekly over a period of 8 weeks. The dose levels utilized were 4, 48, or 100 mg/kg bromoform.
A positive and a negative control group were included in the study design and each contained 20
animals. Mice were sacrificed 24 weeks after the first injection and the frequency of lung tumors
in each test group was compared with vehicle-treated controls. Bromoform produced a
significant increase (p = 0.041) in tumor frequency only at the intermediate dose. U.S. EPA
(1980) concluded that these results were suggestive of carcinogenic activity but were not an
adequate basis for the assessment of cancer risk.
Kurokawa (1987) observed no evidence of carcinogenicity in male or female rats exposed
to microencapsulated bromoform at concentrations of 400, 1600, or 6500 ppm in the diet for 24
months.
Melnick et al. (1998) exposed female B6C3F, mice (10 animals/group) to bromoform in
corn oil via gavage for 3 weeks (5 days/ week). Doses of bromoform used in this study were 0
(vehicle only), 200, or 500 mg/kg-day. A dose-related increase in absolute liver weight and liver
weight/body weight ratio was noted in both tested doses. Neither serum ALT nor serum SDH
activity was significantly elevated at either dose of bromoform. At necropsy, there was no
evidence of hepatocyte hydropic degeneration in animals treated with either dose. BrdU was
administered to the animals during the last 6 days of the study, and hepatocyte labeling index (LI)
analysis was conducted. Only the 500 mg/kg-day dose resulted in marginally significant increase
in hepatocyte proliferation as measured by the LI. Using the Hill equation model, these authors
compared the dose response for liver toxicity (enzyme and LI data) and tumorigenicity (using data
from previously published NTP bioassays) for bromoform. This analysis indicated that the shape
of the dose response as well as the Hill exponents were different for liver toxicity and
tumorigenicity. The authors concluded that these data do not support a causal relationship
between liver toxicity/reparative hyperplasia and tumor development.
De Angelo et al. (2002) evaluated the ability of bromoform administered in drinking water
to induce aberrant crypt foci (ACF), putative early preneoplastic lesions, in the colons of male
F344/N rats. Groups of weanling rats (6 animals/group) were exposed to distilled water, 0.25%
Alkamuls EL-620®, or 1.1 g/L bromoform in 0.25% Alkamuls EL-620 for 13 weeks. A single
intraperitoneal injection of 30 mg/kg azoxymethane (AOM) served as the positive control. Body
weight and water consumption were measured twice during the first week of the study and once
per week thereafter. Colons were collected at study termination, fixed, stained with 0.2%
methylene blue, divided into three equal segments, and examined for ACF. The measured
concentration of bromoform averaged 0.98 ± 0.08 mg/L over the course of the study. When
adjusted for volatilization and adherence to glass, the corrected concentration was 0.77 mg/L.
Water consumption was significantly reduced (30%) in the bromoform exposure group when
compared to the deionized water control. The average daily dose of bromoform was 76 mg/kg-
day as calculated by the study authors. The average terminal body weight of the rats exposed to
bromoform was within 10% of the control values. No ACF were observed in colons from control
animals. ACF were observed in four of six colons from bromoform-exposed animals. The total
number of ACF (26) and number of aberrant crypts per focus (3.71 ±0.36) were significantly
increased relative to the combined deionized water and vehicle controls. Fourteen percent of the
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observed ACF were located in the middle segment of the colon and 86% were located in the distal
(rectal) segment. In comparison, 807 ACF and 4.95 ± 0.25 crypts per focus (mean and standard
error) were observed in the AOM positive control group. Eight percent, 42% and 50% of the
ACF induced by AOM were located in the proximal, middle, and distal segment of the colon,
respectively.
H. Other Key Health Effects
I. Immunotoxicity
a. Bromodichloromethane
Munson et al. (1982) administered bromodichloromethane by gavage to CD-I male and
female mice (8-12/sex/dose) for 14 days at levels of 0, 50, 125, or 250 mg/kg-day.
Bromodichloromethane appeared to affect the humoral immune system, as judged by decreased
antibody-forming (ABF) cells in serum and by decreased hemagglutination titers. These changes
were clearly significant (p < 0.05) at the high dose in both males and females, and decreased ABF
cells were also noted at the mid dose (125 mg/kg-day) in females. This study identified a NOAEL
of 50 mg/kg-day and a LOAEL of 125 mg/kg-day for bromodichloromethane on the basis of
decreased immune function in females Additional information on other endpoints measured in this
study is provided in Section V.B. 1.
French et al. (1999) investigated the immunotoxicity of bromodichloromethane in a series
of four experiments conducted in mice and rats. Immunotoxicity in mice was examined following
exposure via ingestion of drinking water or by gavage. The immunological parameters examined
were antibody response to injected sheep red blood cells and T and B lymphocyte proliferation.
Mitogens used in the proliferation assay were concanavalin A (Con A) or phyto-hemagglutinin-p
(PHA) for T cells and lipopolysaccharide (LPS) for B cells. Female C57BL/6 mice (6 animals per
group) were treated for 14 or 28 days with drinking water containing 0, 0.05, 0.25 or 0.5 g/L
bromodichloromethane. All drinking water (including controls) contained 0.25% Emulphor® to
reduce volatilization of bromodichloromethane. Based on measured water consumption, these
concentrations were estimated by the authors to be equivalent to 0, 10, 37 or 62 mg/kg-day.
There were no significant differences in the number of antibody forming cells, antibody
production, or spleen weights in any treatment group. Likewise, splenic and mesenteric lymph
node cell proliferative responses to T and B cell mitogens were similar in all groups. Continuation
of this study for an additional 2 weeks did not affect any measured parameter. These data identify
a NOAEL of 62 mg/kg-day for short-term exposure.
French et al. (1999) conducted a second experiment in which female C57BL/6 mice were
dosed by gavage with bromodichloromethane in 10% Emulphor® once a day for 16 days.
Treatment groups (6 animals per group) included controls (deionized water or 10% Emulphor®),
50, 125 or 250 mg/kg-day bromodichloromethane. As in the previous experiment, there were no
differences in ABF cells, antibody titers or mitogen-induced proliferation in any treatment groups.
A decrease in spleen weight and spleen-to-weight ratio was observed in the 125 mg/kg-day group
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when compared to the Emulphor® control. However, spleen weights in the Emulphor® control
were significantly higher than those in the deionized water control group, making this finding
difficult to interpret.
French et al. (1999) investigated the immunotoxicity of bromodichloromethane in male
Fisher 344 rats following two different in vivo exposure regimens: ingestion of drinking water
containing bromodichloromethane and gavage. The immunological parameters examined were
antibody response to injected sheep red blood cells and T and B lymphocyte proliferation. The
mitogens used in the proliferation assay were concanavalin A (Con A) or phyto-hemagglutin-p
(PHA) for T cells and S. typhimurium mitogen (STM) for B cells. Six rats per treatment group
were exposed for 26 weeks to drinking water containing 0, 0.07 or 0.7 g/L bromodichloro-
methane and 0.25% Emulphor®. Based on water consumption measurements, these
concentrations were estimated by the authors to be equivalent to average daily doses of 0, 5 or 49
mg/kg-day. There was a significant suppression of Con A-stimulated proliferation of spleen cells
observed in the 49 mg/kg-day dose group. No effect on other immunological parameters was
reported. These data suggest NOAEL and LOAEL values of 5 and 49 mg/kg-day, respectively,
for immunotoxic effects.
French et al. (1999) also examined the effect of short-term exposure to relatively large
doses of bromodichloromethane on immune function. Female F344 rats (6 animals/group)
received gavage doses of deionized water, 10% Emulphor®, or 75, 150, or 300 mg
bromodichloromethane/kg in 10% Emulphor® for 5 days. Surviving high-dose animals had
decreased body, spleen, and thymus weights. Con A and PHA responses were depressed in
spleen cells isolated from high-dose animals. Two of the six rats in the 300 mg/kg-day group died
during the exposure period. The remaining high-dose animals had significantly decreased body,
spleen and thymus weights compared to both control groups. Thymus weight, but not spleen or
body weight, was also decreased in the 150 mg/kg-day group. Con A responses were
significantly depressed in both spleen and mesenteric lymph node (MLN) cells in the 300 mg/kg-
day treatment group. All three (75, 150 and 300 mg/kg-day) dose groups exhibited suppression
of PHA stimulated MLN cells when compared to the vehicle (but not the water) controls. This
discrepancy was due to the fact that Emulphor® alone significantly elevated the proliferative
response to PHA in MLN cells relative to the deionized water group. In contrast to the T cell
responses, there was a significant increase in antibody production and proliferative responses to
STM (B cells) from spleen cells at the highest dose tested (300 mg/kg-day dose group). These
data suggest a marginal NOAEL of 150 mg/kg-day and a LOAEL of 300 mg/kg-day for acute
exposure based on depression of immune response.
b. Dibromochloromethane
Munson et al. (1982) administered dibromochloromethane by gavage to CD-I male and
female mice (8 to 12/sex/dose) for 14 days at levels of 0, 50, 125, or 250 mg/kg-day and
evaluated humoral and cell-mediated immune system functions. Dibromochloromethane
appeared to affect the humoral immune system, as judged by decreased antibody-forming (ABF)
cells in serum and by decreased hemagglutination titers. These changes were significant
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(p < 0.05) at the high dose in both males and females. Decreased ABF cells were also noted at the
mid dose (125 mg/kg-day) in females. This study identified a NOAEL of 50 mg/kg-day and a
LOAEL of 125 mg/kg-day for dibromochloromethane on the basis of decreased immune function
in females. Additional information on this study is provided in Section V.B.2.
c. Bromoform
Munson et al. (1982) administered bromoform (aqueous) by gavage to CD-I male and
female mice (6 to 12/sex/dose) for 14 days at levels of 0, 50, 125, or 250 mg/kg-day. Endpoints
evaluated included humoral immune system function. The authors judged that the humoral
immune system was not significantly affected by bromoform, although a decrease in antibody
forming (ABF) cells was reported for high-dose males. These data suggest a NOAEL of 250
mg/kg-day for effects of bromoform on the immune system. Additional information on this study
is provided in Section V.B.3.
2. Hormonal disruption
No studies or case reports were identified that described hormonal disruption by
dibromochloromethane or bromoform.
Bielmeier et al. (2001) examined the effect of bromodichloromethane on serum
progesterone and luteinizing hormone levels in two experiments conducted as part of an
investigation on full litter resorption (FLR) in F344 rats (see Section V.E.I for additional details).
In the first experiment, rats (7 to 10/treatment group) were administered a single 100 mg/kg dose
by aqueous gavage on gestation day 8 or 9. Hormone levels in samples of tail blood were
determined on GD 9 through 12. FLR was observed in 0, 60 and 100% of the control, GD 8-
dosed, and GD 9-dosed animals, respectively. A marked reduction in progesterone levels was
noted 24 hours after dosing in all rats that resorbed their litters when compared to controls and to
bromodichloromethane-treated animals that retained their litters. The mean progesterone levels in
animals dosed on GD 9 decreased from 137.94 ng/mL ± 11.44 ng/mL to 48.45 ± 23.57 ng/mL
within 24 hours (n = 9). For animals treated on GD 8, the mean progesterone level 24 hours after
bromodichloromethane treatment was 67.01 ± 16.22 ng/mL in animals that resorbed litters (n = 6)
and 127.19 ±14.89 in controls (n=7). The resorbed groups had reduced progesterone levels
comparable to the progesterone levels in non-pregnant animals (n = 2) when assayed three days
after compound administration. In contrast to the effect noted on progesterone levels,
administration of bromodichloromethane had no apparent effect on LH when measured 24 hours
after dosing. However, elevated LH concentrations were observed on GD 11 to 12 in animals
experiencing resorption. LH levels in these groups increased from approximately 0.20 ng/mL on
GD 10 to approximately 0.80 ng/mL on GD 11 and remained elevated through GD 12. In
contrast, LH levels in the controls decreased from 0.31 to 0.14 ng/mL over the same time period.
Bielmeier et al. (2001) performed a second experiment to further characterize the effect of
bromodichloromethane treatment on progesterone and LH levels in pregnant F344 rats. The rats
(8-11/treatment group) were dosed with 0, 75, or 100 mg/kg by aqueous gavage on GD 9. Blood
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samples were collected at 0, 6, 12, and 24 hours after dosing. The incidence of FLR was 0, 64%,
and 90% in the 0, 75, and 100 mg/kg dose groups, respectively. The progesterone levels peaked
in all dose groups (including controls) at 6 hours. At 12 and 24 hours, the progesterone levels in
bromodichloromethane-treated animals that resorbed their litters were progressively reduced.
Progesterone levels in bromodichloromethane-treated animals that retained their litters remained
comparable to levels observed in the control group. No significant differences in LH
concentration were noted among dose groups at any time point.
3. Structure-Activity Relationships
Although the mechanism of brominated trihalomethane toxicity is not known with
certainty, there is abundant evidence to indicate that adverse effects are secondary to metabolism.
Bromine is generally a better leaving group than chlorine, suggesting that bromine substitution
could potentially influence the pathway and rate of trihalomethane metabolism. Multiple studies
(described in Section III.C) indicate that metabolism of chloroform and the brominated
trihalomethanes can occur through one or both of two cytochrome P450-mediated pathways:
reductive metabolism to free radical intermediates or oxidative metabolism to dihalocarbonyls
(Figure 4-1). Although comparative data are limited, there is some evidence to indicate that
chloroform and the brominated trihalomethanes are metabolized to a different extent by these
pathways. Tomasi et al. (1985) examined the reductive metabolism of chloroform,
bromodichloromethane, and bromoform in rats and obtained the following rank order for free
radical formation: bromoform>bromodichloromethane>chloroform. Wolf et al. (1977) reported
that bromoform was more extensively metabolized under anaerobic conditions in vitro than was
chloroform. Gao and Pegram (1992) observed that binding of reactive intermediates to rat
hepatic microsomal lipids and proteins was more than twice as high for bromodichloromethane as
for chloroform when assayed under anaerobic conditions. These results collectively suggest that
reductive metabolism may be a more important metabolic pathway for brominated
trihalomethanes than for chloroform. At present, this apparent difference in metabolism has not
been linked to specific differences in toxicity.
Two mutagenicity studies provide additional information on structure-activity
relationships among the trihalomethanes. Additional details of these studies are presented in
Section V.F. Examination of mutagenicity in a strain of Salmonella typhimurium engineered to
express rat theta-class glutathione-S-transferase (GST) indicated the following order for
mutagenic potency (number of revertants/ppm) of the brominated trihalomethanes:
bromoform ~dibromochlorornethane>brornodichlorornethane (DeMarini et al., 1997). The
potency of the first two compounds was several times greater than that observed for
bromodichloromethane. Analysis of the mutational spectra of the brominated trihalomethanes
indicated that all three compounds have similar mutational spectra (predominately GC—>AT
transitions) and site specificity (middle C of a CCC sequence in target DNA). These observations
suggest that a common reactive intermediate or class of intermediates is likely to mediate the
mutagenicity of these compounds.
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In the second study, Pegram et al.(1997) compared the glutathione »Y-transferase-mediated
mutagenicity of bromodichloromethane and chloroform in a GST+ strain of S. typhimurium (See
section V.F. 1). Revertants were produced in a dose-related manner in the presence of low as well
as high concentrations of bromodichloromethane. In contrast, chloroform induced a doubling of
the number of revertants only at high concentrations. This result provides evidence that bromine
substitution of trihalomethanes confers the capability for GST-catalyzed transformation to
mutagenic intermediates at low substrate concentrations. These data further suggest that
chloroform and the brominated trihalomethanes may induce adverse effects via different modes of
action, and indicate the need for care in extrapolating the characteristics of chloroform
metabolism and toxicity to brominated trihalomethanes.
I. Summary
1. Health Effects of Acute and Short Term Exposure of Animals
Large oral doses of brominated trihalomethanes are lethal to laboratory animals. Reported
acute LD50 values range from 450 to 969 mg/kg for bromodichloromethane, 800 to 1,200 mg/kg
for dibromochloromethane, and 1,388 to 1,550 mg/kg for bromoform. Acute lethality values are
summarized in Table V-l.
Acute oral exposure to sublethal doses of brominated trihalomethanes can also produce
effects on the central nervous system, liver, kidney, and heart. Acute duration studies
investigating endpoints other than death are summarized in Table V-2. Ataxia, anaesthesia,
and/or sedation were noted in mice receiving 500 mg/kg bromodichloromethane, 500 mg/kg
dibromochloromethane, or 1,000 mg/kg bromoform. Renal tubule degeneration, necrosis, and
elevated levels of urinary markers of renal toxicity have been observed in rats receiving 200 to
400 mg/kg bromodichloromethane. Elevated levels of serum markers for hepatotoxicity and
have been observed in rats at doses of bromodichloromethane ranging from approximately 82 to
400 mg/kg-day, and hepatocellular degeneration and necrosis were observed at 400 mg/kg.
Effects on heart contractility were reported in male rats at doses of 333 and 667 mg/kg
dibromochloromethane.
Short term studies of brominated trihalomethanes are summarized in Table V-3.
Short-term exposure of laboratory animals to brominated trihalomethanes has been observed to
cause effects on the liver and kidney. Hepatic effects, including organ weight changes, elevated
serum enzyme levels, and histopathological changes, became evident in mice and/or rats
administered 38 to 250 mg/kg-day bromodichloromethane, 147 to 500 mg/kg-day
dibromochloromethane, or 187 to 289 mg/kg-day bromoform for 14 to 30 days. Kidney effects,
characterized by decreased p-aminohippurate uptake, histopathological changes, and organ weight
changes, became evident in mice and/or rats administered 148 to 300 mg/kg-day
bromodichloromethane, 147 to 500 mg/kg-day dibromochloromethane, or 289 mg/kg-day
bromoform for 14 days. Evidence for decreased immune function was noted at
bromodichloromethane or dibromochloromethane doses of 125 mg/kg-day. Studies examining
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strain differences in response to short-term brominated trihalomethane exposure have not been
reported.
2. Health Effects of Longer-term Exposure of Animals
Subchronic studies of brominated trihalomethanes are summarized in Table V-4. The
predominant effects of subchronic oral exposure occur in the liver and kidney. The effects
produced in these two organs are similar in nature to those described for short-term exposures,
with liver appearing to be the most sensitive target organ for dibromochloromethane and
bromoform exposure. Histopathological changes in the liver were reported in mice and/or rats
administered 200 mg/kg-day bromodichloromethane, 50 to 250 mg/kg-day dibromochloro-
methane, or 50 to 283 mg/kg-day bromoform. Histopathological changes in the kidney were
reported in mice and/or rats administered 100 mg/kg-day bromodichloromethane, or 250 mg/kg-
day dibromochloromethane. Studies examining strain differences in response to subchronic
brominated trihalomethane exposure have not been reported.
Chronic toxicity studies of brominated trihalomethanes are summarized in Table V-5. As
observed for exposure for shorter durations, the predominant effects of chronic oral exposure are
observed in the liver and kidney. Histopathological signs of hepatic toxicity in mice and/or rats
became evident at doses of 6 to 50 mg/kg-day for bromodichloromethane, 40 to 50 mg/kg-day for
dibromochloromethane, and 90 to 152 mg/kg-day for bromoform. Signs of
bromodichloromethane-induced renal toxicity became evident in mice and rats treated with doses
of 25 and 50 mg/kg-day, respectively. Studies examining strain differences in response to chronic
brominated trihalomethane exposure have not been reported.
3. Reproductive and Developmental Effects
Reproductive and developmental studies of brominated trihalomethanes are summarized in
Table V-9. Data on the developmental effects of brominated trihalomethanes suggest that these
chemicals are toxic to the fetus in most cases only at doses that result in maternal toxicity. Signs
of maternal toxicity (decreased body weight, body weight gain and/or changes in organ weight)
were reported in rats administered bromodichloromethane at 25 to 200 mg/kg-day and in rabbits
administered 4.9 to 35.6 mg/kg-day. Signs of maternal toxicity were observed in rats or mice
administered 17 (marginal) to 200 mg/kg-day dibromochloromethane and in mice administered
100 mg/kg-day bromoform. Maternal toxicity was not observed in female rats dosed with up to
200 mg/kg-day of bromoform. Several well-conducted studies on the developmental toxicity of
bromodichloromethane gave negative results at doses up to 116 mg/kg-day in rats and 76 mg/kg-
day in rabbits when administered in drinking water. However, in other studies, slightly decreased
numbers of ossification sites in the hindlimb and forelimb were observed in fetuses of rats
administered 45 mg/kg-day in the drinking water on gestation days 6 to 2land sternebral
aberrations were observed in the offspring of rats administered 200 mg/kg-day by gavage in corn
oil. Reductions in mean pup weight gain and pup weight were observed when the pups were
administered bromodichloromethane in the drinking water at concentrations of 150 ppm and
above (biologically meaningful estimates of intake on a mg/kg-day basis could not be calculated
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for this study). Full litter resorption has been noted in F344 rats, but not Sprague-Dawley rats,
treated with bromodichloromethane at doses of 50 to 100 mg/kg-day during gestation days 6 to
10. Chronic oral exposure to bromodichloromethane resulted in reduced sperm velocities at
doses of 39 mg/kg-day. This response was not accompanied by histopathological changes in any
reproductive tissue examined. Adverse clinical signs and reduced body weight and body weight
gain were observed in parental generation female rats and Fx male and female rats at 150 ppm
(approximately 11.6 to 40.2 mg/kg-day) in a two generation study of bromodichloromethane
administered in drinking water. In the same study, small but statistically significant delays in
sexual maturation occurred in F, males at 50 ppm (approximately 11.6 to 40.2 mg/kg-day) and
F1 females at 450 ppm (approximately 29.5 to 109 mg/kg-day). These delays may have been
secondary to dehydration caused by taste aversion to bromodichloromethane in the drinking
water.
Four of five studies on the reproductive or developmental toxicity of dibromochloro-
methane gave negative results when tested at doses of up to 200 mg/kg-day. In the fifth study,
dibromochloromethane administered at 17 mg/kg-day in a multigenerational study resulted in
reduced day 14 postnatal in one of two F2 generation litters. At 171 mg/kg-day, the mid-dose in
the study, decreased litter size, viability index, lactation index, and postnatal body weight were
observed in some F1 and/or F2 generation. The developmental and reproductive toxicity of
bromoform was examined in two studies.
Bromoform administered to rats at 100 mg/kg-day in corn oil by gavage resulted in a
significant increase in sternebral aberrations in the apparent absence of maternal toxicity. In a
continuous breeding toxicity protocol, gavage doses of 200 mg/kg-day in corn oil resulted
decreased postnatal survival, organ weight changes, and liver histopathology in F1 mice of both
sexes. No effects on fertility or other reproductive endpoints were noted.
4. Mutagenicity and Genotoxicity
In vitro and in vivo studies of the mutagenic and genotoxic potential of
bromodichloromethane, dibromochloromethane, and bromoform have yielded mixed results.
Synthesis of the overall weight of evidence from these studies is complicated by the use of a
variety of testing protocols, different strains of test organisms, different activating systems,
different dose levels, different exposure methods (gas versus liquid), and in some cases, problems
due to evaporation of the test chemical. Overall, a majority of studies yielded more positive
results for bromoform and bromodichloromethane. The genotoxicity and mutagenicity data for
dibromochloromethane are variable. Recent studies in strains of Salmonella engineered to
contain rat theta-class glutathione S-transferase suggest that mutagenicity of the brominated
trihalomethanes may be mediated by glutathione conjugation.
5. Carcinogenicity Studies in Animals
The carcinogenic potential of individual brominated trihalomethanes administered in oil
has been investigated in chronic oral exposure studies in mice and rats. Ingestion of
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bromodichloromethane caused liver tumors in female mice, renal tumors in male mice and in male
and female rats, and tumors of the large intestine in male and female rats. Ingestion of
dibromochloromethane caused liver tumors in male and female mice, and ingestion of bromoform
caused intestinal tumors in male and female rats. Comparison of dose response data for
hepatotoxicity, cell proliferation, and tumorigenesis in female mice suggests that the hepatic
carcinogenicity of brominated trihalomethanes is not a simple consequence of cytotoxicity and
regenerative cell proliferation
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6. Other Key effects
The immunotoxicity of brominated trihalomethanes has been investigated in mice and rats.
Short-term bromodichloromethane exposure resulted in decreased antibody forming cells in
serum, decreased hemagglutinin titers, and/or suppression of Con A-stimulated proliferation of
spleen cells at doses of 125 to 250 mg/kg-day.
No evidence has been reported for hormonal effects following exposure to
dibromochloromethane or bromoform. Studies in pregnant F344 rats detected decreased levels of
progesterone in animals administered 75 or 100 mg/kg bromodichloromethane by aqueous gavage
on gestation day 8 or 9. Increased levels of luteinizing hormone were observed two to three days
after dose administration. Disruption of luteal responsiveness to luteinizing hormone has been
proposed as a possible mode of action by which bromodichloromethane elicits full litter resorption
in F344 rats.
Limited structure-activity data for brominated trihalomethanes and chloroform suggest
that bromination may influence the proportion of compound metabolized via the oxidative and
reductive pathways, with brominated compounds being more extensively metabolized by the
reductive pathway. Additional evidence suggests that a GSH-mediated pathway may play an
important role in metabolism of brominated trihalomethanes, but not chloroform. These data raise
the possibility that brominated trihalomethanes may induce some adverse effects via different
modes of action than chloroform.
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VI. HEALTH EFFECTS IN HUMANS
A. Clinical Case Studies
1. Bromodichloromethane
No clinical reports or short term studies were located on the effects in humans from
ingestion of bromodichloromethane.
2. Dibromochloromethane
No clinical case reports or short term studies were located on the effects in humans from
ingestion of dibromochloromethane.
3. Bromoform
In the past, bromoform was used as a sedative for children with whooping cough. Typical
doses were approximately one drop (about 180 mg), given three to six times/day (Burton-
Fanning, 1901). This dosing usually resulted in mild sedation in children, although a few rare
instances of death or near-death were reported (e.g., Dwelle, 1903; Benson, 1907). These cases
were believed to be due to accidental overdoses. Based on these clinical observations, the
estimated lethal dose for a 10- to 20-kg child is approximately 300 mg/kg, and the LOAEL for
mild sedation is approximately 54 mg/kg-day.
B. Epidemiological Studies
Multiple epidemiological studies have investigated the relationship between exposure to
disinfection by-products in chlorinated drinking water and adverse health effects. These studies
fall into two basic categories: studies of association with cancer (Table VI-1) and studies of
association with adverse pregnancy or birth outcomes or alteration of menstrual function (Table
VI-2). Because the purpose of this document is to isolate the health effects of individual
brominated trihalomethanes, a detailed examination of all available studies on disinfection
byproducts is beyond the scope of this report. Epidemiologic studies published prior to 1994 are
discussed in greater detail in the Drinking Water Criteria on Chlorine (U.S. EPA, 1994a). A
number of recent publications have reviewed the association between chlorination disinfection by-
products and cancer and adverse reproductive or developmental outcomes (e.g., Reif et al., 1996;
Mills et al., 1998; Nieuwenhuijsen et al.,2000; Bove et al., 2002).
A subset of epidemiologic studies has examined possible associations between exposure to
bromodichloromethane and adverse reproductive outcomes. The relationship between exposure
to brominated trihalomethanes and alterations in menstrual function has also been investigated.
These studies are described in greater detail in the sections below.
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Table VI-1 Epidemiological Studies Investigating an Association Between Chlorinated
Drinking Water and Cancer
Reference
Study Description
Observation
Alavanja et al. (1978)
Case control study in seven New York
State counties.
Greater risk of gastrointestinal and
urinary tract cancer mortality, both
sexes, in chlorinated water areas of the
counties.
Cantor et al. (1978)
Ecological study using age-standardized
cancer mortality rates, 1968-1971; and
halomethane levels from U.S. EPA
surveys.
Strongest correlation between bromine-
containing trihalomethanes and bladder
cancer.
Struba (1979)
Case-control study of mortality in North
Carolina, 1975-1978.
Small but significant odds ratios for
rectum, colon and bladder cancers in
rural areas but not in urban areas.
Brenniman et al. (1980)
Case-control study in 70 Illinois
communities, 1973-1976.
Questionnaires sent to water treatment
plants to verily 1963 inventory data on
chlorine levels.
Statistically significant relative risks of
cancer of gall bladder, large intestine,
and total gastrointestinal and urinary
tract in females served by systems with
chlorinated versus nonchlorinated
ground water. Due to many
uncontrolled confounding factors,
authors concluded that chlorination was
not a major factor in the etiology of
gastrointestinal and urinary tract
cancers.
Gottlieb et al. (1981)
Case-control study using mortality data
in Louisiana and estimations of
exposure.
Rectal cancer significantly elevated
with respect to surface or Mississippi
River water consumption.
Young etal. (1981)
Case-control study in Wisconsin, 1972-
1977. Questionnaires sent to waterworks
superintendents on chlorine content.
Colon cancer showed significant
(p<0.05) association with chlorine
intake in all three dosage categories.
Cragle et al. (1985)
Case-control study using colon cancer
cases from seven hospitals in North
Carolina.
Consumption of chlorinated water
strongly associated with colon cancer,
above age 60.
Young et al. (1987)
Case-control study of colon cancer cases
in Wisconsin. Water consumption was
determined by interview, and chloroform
levels by historical records and
measurement.
No association found between
trihalomethane exposure and colon
cancer incidence.
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Table VI-1 (cont.)
Reference
Study Description
Observation
Morris et al. (1992)
Meta-analysis of nine case-control
studies and one cohort study analyzing
cancer and consumption of chlorinated
water or water containing high
chloroform levels.
Statistically significant relative risk of
rectal cancer and bladder cancer in
exposed groups. No colon cancer.
McGheehin et al. (1993)
Population-based case-control study
Association between bladder cancer risk
and exposure to chlorinated water and
trihalomethanes.
King and Marret (1996)
Case-control study conducted by Health
Canada
Increased risk of bladder cancer
associated with total trihalomethane
exposure.
Hildesheim et al.
(1997)
Population-based case-control study of
colon and rectal cancer risk. Iowa,
1986-1989.
Rectal cancer risk associated with
duration of chlorinated water use. No
association of colon cancer risk with
duration of chlorinated water use or
trihalomethane estimates.
Cantor et al. (1998)
Population-based case-control study of
bladder cancer risk. Iowa, 1986-1989.
Positive findings for risk restricted to
men and to current or former smokers.
In men, smoking and exposure to
chlorinated water enhanced the risk of
bladder cancer.
Marcus et al. (1998)
Ecologic study of association between
TTHM in 71 North Carolina public
water supplies and incidence of
histologically confirmed female invasive
breast cancer obtained from cancer
registry data.
TTHM levels not associated with breast
cancer risk when adjusted for potential
confounding factors. Data were
consistent with TTHMs being unrelated
or weakly related to breast cancer risk.
Table VI-2 Epidemiological Studies Investigating an Association Between Chlorinated
Drinking Water and Adverse Pregnancy Outcomes or Altered Menstrual Function
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Table VI-2 (cont.)
Reference
Study Description
Observation
Aschengrau et al. 1989
Hospital-based case-control study of
spontaneous abortion and multiple water
quality parameters in Boston, MA area.
After adjustment for potential
confounders and chemical constituents,
frequency of spontaneous abortion was
increased for consumption of surface
water when compared to use of mixed
surface and ground water (OR 2.2, 95%
C.I. 1.3 - 3.6) The association between
surfance water and increased risk of
spontaneous abortion was not confirmed
by a comparison of chlorinated vs.
chloraminated surface water.
Chloraminated water was used as a
surrogate for low exposure to
disinfection by-products.
Aschengrau et al. 1993
Case-control study of drinking water
quality and occurrence of late adverse
effects among women who delivered
infants during August 1977 - March
1980 in Massachusetts
After adjustment for confounding,
frequency of stillbirths was increased
for women exposed to chlorinated
surface water (OR 2.6, 95% CI 0.9-7.5).
Nuckols et al. (1995)
Cross-sectional study in Colorado of
populations drinking chlorinated and
chloraminated water
No statistically significant effects of
exposure, although odds ratio was
elevated for risk of low birth weight
infants.
Bove et al. (1995)
Cross-sectional study in New Jersey
Relationship between total
trihalomethane levels and "small for
gestational age."
Savitz et al. (1995)
Population based case-control study in
North Carolina
Statistically significant association of
miscarriage with increasing
concentration of TTHM and with the
highest sextile of exposure(OR=2.8,
95% C.I. 1.1, 2.7), but no relationship
with ingested dose or water source.
Small increase in risk of low birth rate.
Gallagher et al. (1998)
Retrospective cohort study of
relationship between THM exposure
during third trimester of pregnancy and
low birthweight, low term birth weight,
and preterm delivery. Colorado birth
certificate data matched to historical
water data based on census block groups.
1990-1993.
Possible association of trihalomethane
concentration in tap water at maternal
residence during third trimester and
risk of term low birth weight deliveries.
Little association with preterm delivery.
Weak association with low birth weight.
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Table VI-2 (cont.)
Reference
Study Description
Observation
Waller etal. (1998)
Prospective study of association between
total and individual THM exposure and
spontaneous abortion. Concurrent THM
data obtained from public water supplies.
Women who drank > 5 glasses/day of
cold tap water containing >75 |ig/L
TTHMs had an adjusted odds ratio of
1.8 for spontaneous abortion. Of
individual THMs, only consumption of
>5 glasses of water containing >18
|ig/L bromodichloromethane (or a
compound co-occurring with
bromodichloromethane) was associated
with spontaneous abortion.
Klotz and Pyrch (1998;
1999)
Case-control study of association
between drinking water contaminants
(including disinfection byproducts) and
neural tube defects. Births with neural
tube defects reported to New Jersey's
Birth Defects Registry in 1993 andl994
were matched against control births
chosen randomly from across the State.
Elevated odds ratios, generally between
1.5 and 2.1, for the association of
neural tube defects with total THMs
(TTHMs). The only statistically
significant results were seen when the
analysis was isolated to those subjects
with the highest THM exposures
(greater than 40 ppb) and was limited to
those subjects with neural tube defects
in which there were no other
malformations (OR = 2.1, 95%CI =
1.1-4.0).
Dodds et al. (1999)
Retrospective cohort study in Nova
Scotia women with singleton births,
1988-1995.
Little association between TTHM level
and fetal weight- or gestational age-
related outcomes. Elevated relative risk
for stillbirths for exposure to >100 |ig/L
TTHM levels during pregnancy. Little
evidence for increased prevalence or
dose-response for congenital
abnormalities with possible exception of
chromosome aberrations for exposure
>100 (ig/L.
Magnus et al. (1999)
Ecologic study in Norway of chlorinated
water consumption and birth defects
observed in births during period 1993-
1995. 1994 data on water quality and
disinfection practice. Water color used
as an indicator for natural organic matter
content.
Among 141,077 births, 1.8% had birth
defects. Adjusted odds ratios (high
color, chlorination vs. low color, no
chlorination) of 1.14 (0.99-1.31) for any
malformation; 1.26 (0.61-2.62) for
neural tube defects; and 1.9 (1.10-3.57)
for urinary tract defects.
Yang et al. (2000)
Study in Taiwan of association between
chlorination of drinking water and low
birth weight.
Examination of 18,025 births showed
no association between consumption of
chlorinated drinking water and low
birth weight.
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Table VI-2 (cont.)
Reference
Study Description
Observation
King et al. (2000)
Population-based retrospective cohort
study in Nova Scotia, Canada to examine
the relationship between TTHM or
individual THMs and risk for stillbirth of
fetuses greater than 500 grams. Study
cohort assembled from a perinatal
database and consisted of 49,756
singleton births that occurred between
1988 and 1995.
Risk doubled for women exposed to a
bromodichloromethane level > 20 |ig/L
when compared to women consuming
concentrations of less than 5 |ig/L
(relative risk = 1.98, 95% confidence
interval of 1.23 - 3.49). When
categories of stillbirth (unexplained
deaths and asphyxia-related deaths)
were examined, relative risk estimates
for asphyxia-related deaths increased by
32% for each 10 |ig/L increase in
bromodichloromethane concentration.
Dodds and King (2001)
Retrospective cohort study conducted
using data from a population-based
perinatal database in Nova Scotia,
Canada and routine water monitoring
data. The cohort consisted of women
who had a singleton birth in Nova Scotia
between 1988 and 1995 and who lived in
an area with a municipal water supply.
Exposure to bromodichloromethane at
concentrations of 20 |ig/L and over was
associated with increased risk of neural
tube defects (adjusted relative risk =
2.5; 95% confidence interval 1.2 to 5.1)
and decreased risk of cardiovascular
anomalies (adjusted relative risk = 0.3;
95% confidence interval 0.2 to 0.7).
No association observed for
bromodichloromethane and cleft
defects.
Waller et al. (2001)
Reanalysis of total trihalomethane
exposure data reported in Waller et al.
(1998).
The study authors reported no apparent
advantage in using a closest-site (vs.
utility-wide) measurement approach for
estimation of exposure to total
trihalomethanes.
Windham et al. (2003)
Prospective study of association between
total and individual THM exposure and
menstrual cycle function. Concurrent
THM data obtained from public water
supplies.
Exposure to dibromochloromethane and
sum of brominated trihalomethanes was
associated with a reductions in length
of the menstrual cycle and follicular
phase of the menstrual cycle, suggesting
possible effects on ovarian function.
Concentrations of >20 |ig/L for
dibromochloromethane and >45 |ig/L
for total brominated trihalomethanes
were associated with reductions in cycle
and follicular phase lengths of
approximately one day. No effect was
noted on length of luteal phase or
duration of menses.
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In their assessment of available data on the available data for reproductive and
developmental effects of disinfection byproducts, Reif et al. (1996) stress that interpretation of
epidemiologic findings for these contaminants are potentially complicated by unmeasured
confounding variables and misclassification errors. Smoking, socioeconomic status, alcohol
consumption, other environmental exposures, and reproductive history are examples of
confounding variables that have the potential to bias estimates of risk in studies of disinfection
byproducts if not measured. Misclassification errors can arise from failure to account for spatial
and temporal variability in contaminant measurement s, migration of study participants, incorrect
assumptions related to water source or use, and use of water treatment data as a surrogate for tap
water concentrations. These factors may result in under- or over-classification of health risks
associated with the consumption of disinfected water. Of greatest concern are variables or errors
which might lead to underestimation of the true public health risks associated with exposure to
tap water containing brominated trihalomethanes. The positive findings in studies of brominated
trihalomethanes thus form a foundation for further studies, but should be interpreted cautiously.
1. Bromodichloromethane
Kramer et al. (1992) conducted a population-based case-control analysis to determine if
exposure to trihalomethanes in drinking water is associated with low birthweight, prematurity, or
intrauterine growth retardation (lower than the 5th percentile of weight for gestational age). A
separate analysis was conducted for each endpoint, using five randomly selected controls for each
affected newborn. Data were collected from Iowa birth certificates from January 1, 1989, to June
30, 1990; the study population was restricted to residents of small towns where all of the drinking
water was derived from a single source (surface water, shallow wells, or deep wells). Exposure
data were based on a 1987 municipal water survey; birth certificate data from 1987 were not used
because data on maternal smoking status first became available in 1989. The study authors
adjusted for maternal age, number of previous children, marital status, education, adequacy of
prenatal care, and maternal smoking. An association was observed between exposure to at least
10 |ig/L bromodichloromethane and intrauterine growth retardation (odds ratio = 1.7). However,
the confidence interval in these cases included 1, indicating that the increases were not statistically
significant.
Waller et al. (1998) conducted a prospective study in pregnant women to examine the
association between trihalomethanes in drinking water and spontaneous abortion (pregnancy loss
at 20 or less completed weeks of gestation). The study participants were recruited from three
facilities of a large managed health care organization which were located in regions of California
that primarily received either mixed, surface, or groundwater. Recruitment occurred when the
women scheduled their first prenatal exam after confirmation of pregnancy. A group of 5,342
subjects completed a telephone interview that obtained information on demographics, previous
pregnancy history, employment status, consumption of tap and bottled water, substance use
(alcohol, tobacco, and caffeine), and other factors. At the time of enrollment in the study, each
woman was at least 18 years of age, at 13 or less weeks of gestation, spoke English or Spanish,
and could provide with certainty the date of her last menstrual period. Following adjustment for
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elective termination of pregnancy, ectopic or molar pregnancies, and multiple gestations, a total of
5,144 pregnancies remained for analysis.
Waller et al. (1998) quantified exposure to trihalomethanes by estimating the subject's
daily tap water intake at 8 weeks gestation. Concentration of total trihalomethanes and any
available data on individual trihalomethanes were obtained directly from the utility supplying
drinking water to a subject's address or zip code. Total trihalomethane levels were calculated by
averaging all measurements taken by the utility supplying a participant's home. Each participant
was assigned a personal exposure classification (high or low) to total trihalomethanes (TTHM)
and individual trihalomethanes (THM) based on the following criteria. A high personal exposure
to TTHM was defined as drinking 5 or more glasses of cold tap water per day and having a
TTHM level of 75 |ig/L or higher. Low personal exposure to TTHM was defined as either 1)
drinking less than 5 glasses of cold tap water per day, 2) having a TTHM level of less than 75
|ig/L, or 3) receiving water from a utility that provided 95% or greater groundwater. Personal
exposures to the individual THMs (bromoform, bromodichloromethane and
dibromochloromethane) were defined in a similar manner, with a high personal exposure being
defined as drinking 5 or more glasses of cold tap water per day with an individual brominated
THM level of 16 |ig/L or higher for bromoform, 18 |ig/L or higher for bromodichloromethane, or
31 |ig/L or higher for dibromochloromethane. Low personal exposures to the individual THMs
were defined as either 1) drinking less than 5 glasses of cold tap water per day, 2) having an
individual THM level below the cutoff, or 3) having a TTHM level less than 72 |ig/L if individual
THM levels were not reported.
The authors found a modest association between consumption of trihalomethane-
containing water and incidence of spontaneous abortion. Increased risk of spontaneous abortion
was noted starting at approximately 75 |ig/L. The adjusted odds ratio (OR) for women who
drank 5 or more glasses of cold tap water per day containing an average TTHM level of 75 |ig/L
or higher during their first trimester was 1.8 (95% confidence interval (C.I.) 1.1, 3.0). An
estimated 18.4% of the study participants were exposed at or above this level. Of the four
individual THMs, only high bromodichloromethane exposure (or exposure to another compound
closely associated with bromodichloromethane) was associated with spontaneous abortion alone
(adjusted OR = 2.0, 95% C.I. 1.2, 3.5) and after adjustment for other THMs (adjusted OR = 3.0,
C.I. 1.4, 6.6). Waller et al. (1998) noted that there was no additive or other effect from
showering or swimming. Therefore, no adjustment was required for these variables.
Misclassification of exposure was identified as the primary limitation of this study. Concentration
levels for most subjects were based on test results for a single day, and thus do not reflect
potential variation in trihalomethane levels over time. In addition, the exposure to THMs from
sources other than ingestion could not be fully characterized.
Because exposure misclassification appeared to be a limitation of the Waller et al. (1998)
study, Waller et al. (2001) reported a reanalysis of exposure data from that study. The objective
of the new analysis was to examine how use of alternative methods for estimation of exposure
would affect associations between TTHM exposure and risk of spontaneous abortion. This
reanalysis did not address dose-response relationships between individual brominated
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trihalomethanes and occurrence of spontaneous abortion. Two exposure analyses were tested.
The first method used the utility-wide average concentration (the metric used in Waller et al.,
1998). The second method used THM measurements taken from the water system sampling site
nearest the subject's home. For each method, the authors performed 1) an unweighted analysis;
2) an analysis weighted by a factor intended to reduce exposure misclassification; and 3) an
analysis within a subset of the cohort that possibly had less exposure misclassification than the
entire cohort. The weighted and subset analyses were performed in an effort to reduce exposure
misclassification. The utility-wide average method estimated the concentration of total
trihalomethanes by averaging all distribution measurements taken by the subject's utility during
the first trimester of pregnancy. In contrast to the method used in Waller et al. (1998) the time
interval was not expanded in order to capture a measurement and thus reduce missing data in
cases where no measurements were available in the first trimester. The closest-site method took
the average of all measurements taken during the first trimester of pregnancy at the water
distribution site nearest to the subject's home. For the weighted analyses, the utility-wide
approach used the variance of the utility-wide average as a proxy for accuracy. The closest-site
approach used TTHM measurements taken from the water system sampling site nearest the
subject's home and adjusted for distance between the subject's home and the sampling site. For
the subset approach, analyses were restricted to groups of women for whom the exposure
assessment was likely to be more accurate. Subset analyses using the utility-wide average TTHM
concentration included women whose utility measurements were all within 20 |ig/L of each other
and women served by groundwater utilities. Subset analyses using the closest-site average
concentrations used women who lived within 0.5 miles of the utility sampling site and all women
served by groundwater utilities. An ingestion metric was calculated using individual daily cold tap
water intake at eight weeks gestation as determined in Waller et al. (1998). A categorical
ingestion exposure metric was created using the first trimester THM concentration dichotomized
at 75 |ig/L and cold tap water ingestion dichotomized at 5 glasses per day. Ingestion was also
estimated by multiplying the TTHM concentration by cold tap water consumption. A metric to
capture exposure to trihalomethanes during showering was created by multiplying THM
concentration by typical shower duration and the frequency of showering.
Use of the utility-wide approach generally resulted in odds ratios equivalent to or slightly
higher than the closest-site approach. Odds ratios obtained using the utility-wide average method
for estimating TTHM (but not the closest-site method) became progressively stronger in the
weighted and subset analyses. The study authors reported a positive, monotonic dose-response
relationship between spontaneous abortion rate and an exposure metric incorporating TTHM and
personal ingestion. Rates of spontaneous abortion obtained using this approach ranged from
8.3% to 13.7% (unweighted); 7.9% to 16.6 % (variance weighted); and 6.6% to 16.6% (low-
variance subset). The study authors noted that a major limitation of this reanalysis is the lack of a
"gold standard" with which to compare the estimated TTHM ingestion of subjects in the study.
In the absence of such a standard, it is not possible to determine whether the reanalysis actually
reduced exposure misclassification. The conclusions reached by the study authors were 1) there
was no advantage in using the closest-site method over the utility-wide method for exposure
analysis; 2) use of variance-based weighting factors and subset analyses is defensible and resulted
in some increases of odds ratio, but resulting loss of sample size may limit the utility of these
techniques; and 3) use of a variety of exposure assessment techniques may lessen the impact of
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bias resulting from utility-specific factors such as inconsistencies in sampling density or
unrecognized contamination problems.
The reanalysis conducted by Waller et al. (2001) identified evidence for differential
misclassification in the prior analysis of a ground water predominate area ("Zone A") reported in
Waller et al. (1998). The effect of this misclassification was to bias the original estimate of the
relationship between TTHM ingestion and spontaneous abortion away from the null. Over 400 of
the women in the study cohort resided in Zone A, an area within a large mixed-source utility that
received predominately groundwater. Zone A was not sampled for THMs during the study
period. Because other areas within the utility frequently had high TTHM concentrations, use of a
utility-wide approach for estimating TTHM concentration probably resulted in an overestimation
of exposure for Zone A residents. An investigation by the study authors revealed that although
the spontaneous abortion rate of women in Zone A was low overall, women who drank at least 5
glasses of water per day had a spontaneous abortion rate of 14.6%. The reason for the high
spontaneous abortion rate among women consuming large amounts of Zone A water was unclear,
but was reported to be consistent with other epidemiological studies that found high rates of
spontaneous abortion among women ingesting large amounts of unchlorinated water in Region 1
of the original study. Exclusion of Zone A residents or recoding them to a level determined by
later testing within the zone decreased the adjusted OR for high exposure to TTHM (TTHM >75
|ig/L and intake > 5 glasses per day) to 1.5 (95% C.I. 0.8, 2.8) as compared to the adjusted OR
of 1.8 (95% C.I. 1.1, 3.0) identified in the original analysis (Waller et al., 1998). The impact of
this finding on the adjusted OR calculated for individual brominated THMs is currently unknown,
but is expected to be addressed in a future publication by Waller et al.
King et al. (2000) conducted a population-based retrospective cohort study to examine the
relationship between TTHM or individual THMs and risk for stillbirth of fetuses greater than 500
grams. The study cohort was assembled from a perinatal database in Nova Scotia, Canada and
consisted of 49,756 singleton births that occurred between 1988 and 1995. Exposure was
assigned by relating the mother's residence at the time of delivery to the levels of total and
individual THMs measured in public water supplies. Maternal age, parity, smoking during
pregnancy, infant's sex, and neighborhood family income were evaluated as potential cofounders.
Relative risks were adjusted for smoking and maternal age. Exposure to TTHMs, chloroform,
and bromodichloromethane were associated with increased risk of stillbirth. Analysis of the
results suggested that exposure to bromodichloromethane was a stronger predictor of risk than
exposure to chloroform. Risk doubled for women exposed to a bromodichloromethane level of
greater than or equal to 20 |ig/L when compared to women consuming concentrations of less than
5 |ig/L (relative risk = 1.98, 95% C.I. 1.23, 3.49). When categories of stillbirth (unexplained
deaths and asphyxia-related deaths) were examined, relative risk estimates for asphyxia-related
deaths increased by 32% for each 10 |ig/L increase in concentration of bromodichloromethane.
As noted by the authors, a potential limitation of this study was misclassification of exposure as a
result of mobility within the study population (estimated to affect less than 10% of study
subjects). This study did not examine early fetal death (e.g. spontaneous abortion) because the
perinatal database employed in this investigation contained information only on fetuses that
weighed 500 grams or more.
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Dodds and King (2001) conducted a retrospective cohort study of singleton births among
49,842 residents of Nova Scotia, Canada between 1988 and 1995 to assess the relationship
between exposure to THMs and birth defects. Of the brominated THMs, only
bromodichloromethane was examined. Information on exposure concentrations consisted of
routine water monitoring data obtained from the Nova Scotia Department of the Environment and
reflected samples collected from within the water distribution system. The birth defects examined
had previously been reported in other epidemiological studies, and included neural tube defects,
cardiovascular defects, cleft defects, and chromosomal abnormalities. The perinatal information
used in the study was obtained from the Nova Scotia Atlee perinatal database. This database
contains information abstracted from medical records and includes infant diagnoses among
stillborn and live born infants up to the time of discharge from the hospital after birth. In addition,
information on prenatally diagnosed congential anomalies was obtained from pregnancy
terminations. Inclusion of this data was deemed important because, in Nova Scotia,
approximately 80% of neural tube defects are detected antenatally and the pregnancy is electively
terminated. Exposure windows were selected to target the period before or during gestation
when exposure to a potential developmental toxicant or mutagen might have the most profound
effect on a particular developmental or genotoxic endpoint. The selected windows were as
follows: average bromoform concentrations from one month and one month after were used for
analysis of neural tube defects; concentrations during the first two months of pregnancy were used
for analysis of cardiac defects and cleft defects; and the average concentrations three months
before pregnancy were used for the analysis of chromosomal abnormalities. Estimates of relative
risks and 95% confidence intervals were obtained from Poisson regression models. Maternal age,
parity, maternal smoking, and neighborhood family income were assessed as potential
confounders. The categories used for bromodichloromethane concentration were <5 |ig/L; 5-9
|ig/L; 10-19 |ig/L; and > 20 |ig/L.
Exposure to bromodichloromethane at concentrations > 20 |ig/L was associated with
increased risk of neural tube defects (adjusted relative risk 2.5; 95% C.I. 1.2, 5.1) when adjusted
for maternal age and income level. However, there was no evidence of a dose-response trend
with increasing concentration of bromodichloromethane. In addition, the study authors noted that
this point estimate was "fairly unstable" as a result of the low number of cases (n=10) in the >20
|ig/L exposure category. For cardiac defects, a significant reduction in risk was associated with
exposure to concentrations of >20 |ig/L (relative risk 0.3; 95% C.I. 0.2, 0.7) and there was a
trend of decreasing risk with increasing exposure. The study authors considered it unlikely that
exposure above >20 |ig/L was actually protective. They suggested that this may be a chance
finding or a reflection of a negative association of bromodichloromethane with other disinfection
by-products in this region which may increase cardiac risks. There was no apparent trend or
significant association for exposure to bromodichloromethane and occurrence of cleft defects or
chromosomal aberrations.
2. Dibromochloromethane
Windham et al. (2003) examined menstrual cycle characteristics in relation to the presence
of brominated trihalomethanes in tap water in a prospective study of women living in Northern
California. Data were also reported for chloroform and TTHM. The relationships examined
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included: 1) cycle characteristics and concentration of individual THMs, TTHM, and total
brominated THMs in tap water; 2) cycle characteristics and estimated water consumption
(TTHM); and 3) cycle characteristics and duration of showering (TTHM). The target population
was married women of reproductive age (18-39 years old) who were members of the Kaiser
Permanente Medical Care Program. Participants in the study were enlisted between May 1990
and June 1991. Participants were selected from among nearly 6500 women using a short
screening interview to identify women who were more likely to become pregnant (i.e., those who
reported a menstrual period within the last six weeks, were not surgically sterilized, did not use
birth control pills or IUDs, and were non-contracepting for less than 3 months). Out of 1092
eligible women, a total of 403 women collected first morning urine samples daily for 2-9
menstrual cycles (average 5.6 cycles) for measurement of steroid metabolites. These
measurements were used to estimate menstrual parameters such as cycle and phase length. Cycle
length was calculated from the first day of menses to the day before onset of the next menses.
When the available data permitted, the cycle was divided into the follicular phase (first day of
menses through estimated day of ovulation) and the subsequent luteal phase. Between 1424 and
1714 cycles were available for evaluation of each parameter. Information on water consumption
(as unheated tap water or drinks made from unheated tap water, drinks made from heated tap
water, and bottled water) and other variables (age, race, education, employment, income,
pregnancy history, exercise type and frequency, smoking, alcohol and caffeine consumption) was
collected in a baseline telephone interview prior to urine collection. Information on the number of
showers taken at home per week and their duration was also collected. Showering was examined
as minutes per week and by combining the duration and cycle-specific THM level to create
combinations of high and low exposure. The participants filled out a daily diary during the urine
collection phase and recorded vaginal bleeding as number of pads or tampons. Exposure to
THMs was estimated from quarterly utility monitoring data and information on drinking water
and other tap water use collected during the baseline interview. A 90-day exposure time period
was assigned for each cycle because THM monitoring was conducted by the utilities on a
quarterly (i.e., about 90 days) basis. A period of 60 days before and 30 days after each cycle start
date was selected for the 90-day window. Cycle-specific exposures to TTHM and individual
THMs were calculated by averaging all THM measurements taken by a participant's utility at
various points in the distribution system (i.e., the "utility-wide average" described by Waller et al.,
1998, 2001) during the that 90-day period. Because the brominated THMs were highly
correlated and thus difficult to examine independently, the study authors also examined the sum of
the levels of the three brominated compounds. Exposure levels for the brominated THMs were
examined as categorical variables. For TTHM, ingestion metrics were calculated for unheated tap
water and the sum of heated and unheated tap water. Statistical analyses were conducted using
the menstrual cycle as the unit of observation. Menstrual parameters were analyzed as continuous
or categorical variables in relation to categorical exposure indices and the methods used
accounted for repeated measures. Numerous covariates reflecting demographic, reproductive
history, and lifestyle factors were examined in relation to categorical trihalomethane levels and
ingestion to identify potential confounders. Age, pregnancy history, body mass index, caffeine
consumption, and alcohol consumption, as well as race and smoking, were included as variables in
all adjusted models.
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Increasing levels of individual brominated compounds or total brominated THMs were
associated with significantly shorter cycles when examined by quartile (Table VII-3). Similar
decrements were observed in follicular, but not luteal, phase length. Dose-response patterns were
evident for both individual and total brominated THMs. The strongest association for an
individual compound was observed for dibromochloromethane, with adjusted decrements of 1.2
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Table VI-3 Means and Adjusted Differences in Menstrual Cycle and Follicular Phase
Length by Quartile of Individual and Summed Brominated Trihalomethanes
Compound
Quartile of Exposure3
lb
2-3
4
Mean in days
(SE)
Adjusted Difference0
(95% CI)
Adjusted Difference
(95% CI)
Cycle Length
Bromodichloromethane
29.8 (0.30)
-0.59 (-1.2, -0.02)
-0.74 (-1.5, -0.02)
Dibromochloromethane
30.0 (0.33)
-0.69 (-1.4, -0.02)
-1.2 (-2.0, -0.38)
Bromoform
29.7 (0.26)
-0.42 (-0.96, 0.13)
-0.79 (-1.4, -0.14)
Sum of Brominated
Compounds
30.0 (0.34)
-0.72 (-1.4, -0.04)
-1.2 (-2.0, -0.40)
Follicular Phase
Bromodichloromethane
17.0 (0.31)
-0.54 (-1.1, 0.01)
-0.80 (-1.5, -0.08)
Dibromochloromethane
17.1 (0.34)
-0.62 (-1.3, 0.05)
-1.1 (-1.9, -0.25)
Bromoform
16.9 (0.27)
-0.30 (-0.83, 0.23)
-0.78 (-1.4, -0.14)
Sum of Brominated
Compounds
17.2 (0.35)
-0.66 (-1.3, 0.02)
-1.1 (-1.9, -0.29)
Source: Windham et al. (2003)
3 Top quartiles for bromodichloromethane, dibromochloromethane, bromoform, and the summed
brominated compounds are >16, >20, >12, and >45 (ig/L, respectively.
b Reference group; the mean provided is unadjusted with standard error (SE)
c Adjusted for age, race, body mass index, income, pregnancy history, caffeine and alcohol
consumption, and smoking.
days (95% C.I. -2.0, -0.38) for mean cycle length and 1.1 days (95% C.I. -1.9, -0.25) for mean
follicular phase length at the highest quartile (>20 |ig/L). In comparison, a clear association with
reduced cycle length was not observed for chloroform (difference -0.3 days; 95% C.I. -1.0, 0.40),
even at the highest quartile (>17 |ig/L). Menses length was slightly increased at the highest
quartile for bromodichloromethane exposure. For categorical variables, the odds of having a long
cycle (Adjusted Odds Ratio, AOR 0.55; 95% C.I. 0.28, 1.08) or long follicular phase (AOR 0.26;
95% C.I. 0.12, 0.59) were significantly reduced at the highest quartile for summed brominated
THM concentration (>45 |ig/L). These data suggest that brominated THMs or other disinfection
by-products that co-occur with brominated THMs may affect ovarian function.
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Windham et al. (2003) also examined relationships between TTHM (brominated
compounds plus chloroform) exposure and menstrual parameters. A monotonic decrease in mean
cycle length was observed with increasing TTHM exposure category. At concentrations greater
than 60 |ig/L, the adjusted decrement was 1.1 day (95% C.I. -1.8, -0.40) when compared to
concentrations less than 40 |ig/L. The decrease in follicular phase length was similar (-0.94 day;
95% C.I. -1.6, -0.24). Cycles with TTHM concentrations above the current MCL of 80|ig/L
were appeared to be shorter by about one day (0.99 days; 95% C.I. -2.2, 0.17). A unit decrement
in mean cycle and follicular phase length of 0.18 days per 10 |ig/L increase in total trihalomethane
concentration (95% C.I. -0.29, -0.07) was determined when the cycle-specific TTHM level was
examined as a continuous variable. When ingestion patterns were examined, mean cycle and
phase lengths showed little variation in relation to consumption of unheated tap water at home.
In contrast, increased consumption of heated tap water was associated with significantly
decreased cycle and follicular phase lengths. The observed decrements were greater than one day
with daily consumption of three or more drinks made from heated tap water. These decrements
were reduced by adjustment, particularly when caffeine was included in the model; the adjusted
decrement in cycle length was 0.68 days (95% C.I. -2.1, 0.72). A non-monotonic relationship
was observed for mean cycle length and an ingestion metric combining TTHM concentration and
consumption of unheated tap water, with the highest category (>60 |ig/L) showing a decrement of
0.4 days and the third category (>40-60 |ig/L) showing a decrement of one day. Use of an
ingestion metric based on total home tap water consumption (i.e., heated and unheated tap water)
revealed a more consistent pattern of reduced cycle length, with adjusted decrements of greater
than one day for cycle (-1.1 day; 95% C.I. -2.2, -0.06) and follicular phase (-1.1 day; 95% C.I. -
2.2, 0.03) lengths. Examination of time spent showering did not reveal additional risks with
longer showers. The unadjusted mean cycle length varied little by time spent showering.
Following adjustment, there was a tendency toward decreased length with any category of
showering above 35 minutes/week. This relationship was stronger for follicular phase duration
than for cycle length. For example, the adjusted mean decrements at the longest duration (> 105
minutes) were -0.68 days (95% C.I. -2.0, 0.68) for cycle length and -1.2 days (95% C.I. -2.6,
0.26). However, the confidence intervals were wide for all duration categories and a clear dose
response pattern (i.e., shorter lengths at higher durations) was not evident.
The study authors noted several strengths and potential limitations of this study.
Strengths include the use of a prospective study design, use of biologic measures to determine
menstrual parameters, and consideration of many potential confounders. Potential limitations
include exposure misclassification, use of a study sample that is not representative of the general
population, lack of information on other sources of exposure such as washing, cooking and
cleaning, as well as exposures outside the home. There are two observations in this study that
might suggest involvement of compounds other than THMs in the reduction of cycle length.
First, the more consistent association of decreased cycle length reported for heated compared to
unheated tap water is unexpected if THMs alone are the causative agent. This is because THMs
volatilize from heated water and exposure to these compounds should therefore be lower for
heated tap water, unless the volatilized compound is inhaled. Second, examination of time spent
showering did not reveal additional risks with longer showers. This is also counter to the
expected trend, as elevated blood levels of THMs have been documented after showering (Backer
et al., 2000; Lynberg et al., 2001) as a result of dermal and inhalation exposure. However,
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information on shower duration was collected by interview and the reported lengths may not have
accurately reflected actual shower duration.
Because the study by Windham et al. (2003) is the first to examine changes in menstrual
cycle function in relation to tap water exposure, there are no supporting data on the association of
disinfection by-products other than the THMs with changes in menstrual cycle function.
Although this study suggests that disinfection by-products may have effects on ovarian function,
no conclusions can be drawn regarding the identity of the compounds responsible for these
effects. The relationships reported in this study underscore the need for additional research on the
effects of drinking water disinfection by-products on menstrual cycle function.
3. Bromoform
Epidemiological studies identifying adverse health effects specifically associated with
exposure to bromoform were not identified.
C. High Risk Populations
High risk (or susceptible) populations are those which experience more adverse effects at
comparable levels of exposure, which experience adverse effects at lower exposure levels than the
general population, or which experience a higher than average exposure because they live or work
in settings with elevated environmental concentrations of the chemical of interest. The enhanced
response of these susceptible subpopulations may result from intrinsic or extrinsic factors.
Factors that may be important include, but are not limited to: impaired function of detoxification,
excretory, or compensatory processes that protect against or reduce toxicity; differences in
physiological protective mechanisms; genetic differences in metabolism; developmental stage;
health status; gender; or age of the individual. For brominated trihalomethanes, high risk
populations may potentially include those with elevated levels of CYP2E1 (via exposure to
inducing substances or because of altered physiological or health states) or elevated levels of
glutathione-»Y-transferase theta. These factors are discussed in greater detail in Section VII.D.3 of
this document.
A growing body of scientific evidence indicates that children may suffer disproportionately
from some environmental health risks. These risks may arise because the neurological,
immunological, and digestive systems of children are still developing (U.S. EPA, (1998a). In
addition, children may incur greater exposure because they eat more food, consume more fluids,
and breathe more air in proportion to their body weight when compared to adults (U.S. EPA,
1998a). Factors contributing to potentially greater risk in children are discussed in Section
VII.D.2 of this document.
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D. Summary
Limited human health data are available for the brominated trihalomethanes. In the past,
bromoform was used as a sedative for children with whooping cough. Doses of 50 to 100 mg/kg-
day usually produced sedation without apparent adverse effects. Some rare instances of death or
near-death were reported, although these cases were generally attributed to accidental overdoses.
No human toxicological data were available for bromodichloromethane or
dibromochloromethane.
Numerous epidemiological studies have examined the association between water
chlorination and increased cancer mortality rates. None of these studies has examined the
association between cancer and exposure to any individual brominated trihalomethane. Recent
studies have examined the association of chlorinated water use with various pregnancy outcomes,
including low birth weight, premature birth, intrauterine growth retardation, spontaneous
abortion, stillbirth , and birth defects. An association has been reported for exposure to
bromodichloromethane (or a closely associated compound) and a moderately increased risk of
spontaneous abortion during the first trimester. A confirmation of this finding is pending
reanalysis of the original data to correct a differential misclassification error identified in a
subsequent analysis of the study data. An association has also been reported for exposure to
bromodichloromethane (or a closely associated compound) 1) and stillbirth of fetuses weighing
more than 500 g and 2) increased risk of neural tube defects in women exposed to >20 |ig/L of
bromodichloromethane prior to conception through the first month of pregnancy. An association
has been reported for total brominated trihalomethanes and reduced menstrual cycle and follicular
phase length in women of child-bearing age. Among the individual brominated trihalomethanes,
dibromochloromethane displayed the strongest association with altered menstrual function. It is
not possible to directly conclude from these studies that bromodichloromethane and
dibromochloromethane are developmental toxicants in humans, because chlorinated water
contains many disinfection by-products. Nevertheless, these studies raise significant concern for
possible human health effects. The methodology used to estimate exposure to brominated
trihalomethanes in tap water has been examined with the goal of refining estimates of intake of
these compounds in epidemiological studies.
For the brominated trihalomethanes, populations at high risk may potentially include those
with elevated levels of CYP2E1 (via exposure to inducing substances or because of altered
physiological or health states) or elevated levels of glutathione-»Y-transferase theta. In addition,
users of hot tubs and swimming pools may experience additional exposure to brominated
trihalomethanes.
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VII. MECHANISM OF TOXICITY
A. Role of Metabolism
The toxicity of the brominated trihalomethanes is related to their metabolism. This
conclusion is based largely on the observation that liver and kidney, the chief target tissues for
these compounds, are also the primary sites of their metabolism. In addition, treatments which
increase or decrease metabolism also tend to increase or decrease trihalomethane-induced toxicity
in parallel. Pankow et al. (1997), for example, examined the relationship between metabolism of
dibromochloromethane and hepatotoxicity. Serum leucine aminopeptidase (LAP) activity (an
indicator of hepatotoxicity) increased in a dose-dependant fashion with any treatment that
increased the metabolism of dibromochloromethane (e.g. pretreatment with isoniazid or m-
xylene).
B. Biochemical Basis of Toxicity
The precise biochemical mechanisms which link brominated trihalomethane metabolism to
toxicity are not certain, but many researchers have proposed that toxicity results from the
production of reactive intermediates. These reactive intermediates are believed to form covalent
adducts with various cellular molecules and to impair the function of those molecules resulting
cell injury. Reactive intermediates may arise from the oxidative (dihalocarbonyls) or the reductive
(free radicals) pathways of metabolism as discussed in Section III.C. Support for this mode of
action has been obtained from in vitro studies of bromodichloromethane. Under both aerobic and
anoxic conditions, bromodichloromethane is metabolized to intermediates that covalently bind to
rat microsome proteins and lipids. Direct evidence showing a relationship between the levels of
covalent binding intermediates generated by the oxidative or reductive pathways and the extent of
toxicity is not currently available for brominated trihalomethanes.
Free radical generation by the reductive pathway for brominated trihalomethane
metabolism may result in lipid peroxidation. Although evidence demonstrating that lipid
peroxidation actually accounts for the observed cellular toxicity associated with brominated
trihalomethanes is lacking, at least one study has established that lipid peroxidation does occur in
conjunction with brominated trihalomethane metabolism. De Groot and Noll (1989) reported that
all three brominated trihalomethanes induced lipid peroxidation in rat liver microsomes in vitro,
and that this was maximal at low oxygen levels (between 1 and 10 mm Hg of 02). The authors
interpreted these data to support the concept that lipid peroxidation is caused by free radical
metabolites generated by the reductive metabolism of trihalomethanes.
Glutathione has been implicated in both defense against toxicity induced by brominated
trihalomethanes and in generation of mutagenic metabolites. Gao et al. (1996) examined the
effect of glutathione on the toxicity of bromodichloromethane in vivo and in vitro. Depletion of
glutathione by pretreatment of male F344 rats with the glutathione synthesis inhibitor buthionine
sulfoximine increased the hepatotoxicity of a single gavage dose of 400 mg/kg
bromodichloromethane administered in 10% Emulphor®. Biochemical indicators of
hepatotoxicity (e.g. AST, ALT, LDH) were increased approximately 11-fold and the severity of
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morphological changes (centrilobular vacuolar degeneration and hepatocellular necrosis) was
greater in the glutathione-depleted animals. Serum and urinary markers of renal damage were
also significantly increased by glutathione depletion. Renal tubule necrosis was observed only in
the glutathione-depleted group. Overall, glutathione depletion appeared to enhance hepatotoxicity
more than nephrotoxicity, an effect that was attributed to organ-specific differences in
bromodichloromethane metabolism. The addition of glutathione to reaction mixtures of rat
hepatic or renal microsomal fraction and radiolabeled bromodichloromethane resulted in 92% and
20% reductions in protein binding to bromodichloromethane, respectively. The difference in
response to glutathione addition was interpreted as evidence for existence of different metabolic
pathways in liver and kidney. Bromodichloromethane binding to lipid in liver microsomes under
anaerobic conditions was decreased in the presence of glutathione, suggesting that glutathione can
react with the dihalomethyl radical.
In contrast to the apparent role protective role of glutathione described above, studies in
strains of S. typhimurium engineered to express rat theta class glutathione S-transferase suggest
that conjugation with glutathione leads to formation of mutagenic metabolites (Pegram et al.,
1997; DeMarini et al., 1997). These studies are described in greater detail in Section V.F.I.
Proposed pathways for generation of the mutagenic species are outlined in Figure V-2, also
located in Section V.F.I. Briefly, similar mutational specificity, site specificity, and mutation
spectra for the three brominated trihalomethanes support the conclusion that they are activated by
one or more common pathways. In contrast, the data do not support a glutathione S-transferase
mediated pathway for the structurally-related trihalomethane chloroform. This finding suggests
that chloroform and the brominated trihalomethanes may in some instances be metabolized by
different pathways.
C. Mode of Action of Carcinogenesis
Administration of individual brominated trihalomethanes has been associated with
formation of liver tumors (bromodichloromethane, dibromochloromethane), kidney tumors
(bromodichloromethane), and tumors of the large intestine (bromodichloromethane, bromoform)
in some experimental animals. The mode of action by which brominated trihalomethanes induce
tumors in laboratory animals is not known. However, two general modes of action have been
proposed: 1) formation of DNA adducts resulting from interaction with one or more classes of
reactive metabolites and 2) production of cytotoxicity coupled with regenerative hyperplasia.
The production of reactive metabolites from trihalomethanes is well-established. Classes
of reactive metabolites produced include dihalocarbonyls produced by oxidative metabolism and
and dihalomethyl radicals produced by reductive metabolism. Additional evidence suggests that
reactive species can be also formed via glutathione conjugation (DeMarini et al., 1997; Pegram et
al., 1997). Detection of adduct formation and consistent evidence of DNA reactivity in standard
assays are two lines of experimental evidence that would strongly support the adduct formation
hypothesis. At present there are no in vivo data available on DNA adducts resulting from
metabolism of brominated trihalomethanes. DNA reactivity can be inferred from test results of
mutagenic and genotoxic potential. As noted previously (U.S. EPA, 1994b), synthesis of the
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overall weight of evidence for genotoxicity of individual brominated trihalomethanes is
complicated by the use of a variety of testing protocols, different strains of test organisms,
different activating systems, different dose levels, different exposure methods (gas versus liquid),
and in some cases, problems due to evaporation of the test chemical. Overall, a majority of
studies yielded more positive results for bromoform and bromodichloromethane, and evidence of
mutagenicity is considered adequate for these chemicals. Study results for the mutagenicity of
dibromochloromethane are mixed, and the overall evidence for mutagenicity of this chemical is
judged to be inconclusive (U.S. EPA, 1994b).
Alternatively, the induction of tumors by individual brominated trihalomethanes could
involve an epigenetic mode of action. Induction of tumors in animal studies has been noted to
occur primarily at sites where cytotoxicity was observed (i.e., liver and kidney), and there appears
to be a correlation between hepatotoxicity and liver tumorigenicity of brominated trihalomethanes
in mice (bromodichloromethane > dibromochloromethane > bromoform) (U.S. EPA, 1994b).
This raises the possibility that regenerative hyperplasia caused by the cytotoxic effects of
brominated trihalomethanes may contribute to the tumorigenic potential of these chemicals. A
brief review of studies that have evaluated regenerative hyperplasia following exposure to
brominated trihalomethanes is provided below.
A number of studies have measured cell proliferation in liver and/or kidney following
exposure to brominated trihalomethanes. Miyagawa et al. (1995) observed evidence for the
induction of hepatocyte proliferation in male B6C3F, mice following a single oral gavage dose of
dibromochloromethane in corn oil at the maximum tolerated dose (MTD) or at one half the MTD
(200 or 400 mg/kg). Proliferation was assessed by incorporation of [3H]- thymidine using the in
vivo-in vitro replicative DNA synthesis assay at 24, 39, and 48 hours postexposure.
Potter et al. (1996) investigated cell proliferation in the kidney of male F344 rats. Test
animals received 0.75 or 1.5 mmol/kg of bromodichloromethane in 4% Emulphor® by gavage for
1, 3, or 7 days. The administered doses corresponded to 123 or 246 mg/kg-day for
bromodichloromethane, 156 or 312 mg/kg-day for dibromochloromethane, and 190 or 379
mg/kg-day for bromoform. Cell proliferation in the kidney was assessed in vivo by [3H]-
thymidine incorporation. No statistically significant effect of bromodichloromethane on tubular
cell proliferation was observed following exposures of up to 7 days, although high labeling levels
were observed in 3 of 4 rats at the 246 mg/kg-day dose of bromodichloromethane.
NTP (1998) evaluated cell proliferation in the kidney and liver of Sprague-Dawley rats as
part of a short-term reproductive and developmental toxicity screen of bromodichloromethane.
The compound was administered in drinking water for 35 days. Groups of male and female rats
were exposed to drinking water concentrations of 0, 100, 700 and 1300 ppm
bromodichloromethane using the study design described in Table V-6 (Section V.D. 1). BrdU
labeling index (LI) was unchanged in the livers and kidneys of Group B males at doses up to 69
mg/kg-day. A small but statistically significant increase in the LI was noted in the livers and
kidneys of Group C females in the 1300 ppm dose group (126 mg/kg-day). The study authors
noted that the result in females was probably biologically insignificant.
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Melnick et al. (1998) exposed female B6C3F, mice (10/dose) to bromodichloromethane,
dibromochloromethane, or bromoform in corn oil via gavage for 3 weeks (5 days/week). BrdU
was administered to the animals during the last 6 days of the study, and hepatocyte labeling index
(LI) analysis was conducted. Time-adjusted doses of 107, 336, and 357 mg/kg of
bromodichloromethane, dibromochloromethane, and bromoform, respectively, resulted in
significantly elevated hepatocyte proliferation as measured by the LI. The authors compared the
dose response for liver toxicity (including hepatic enzyme activity and LI data) and tumorigenicity
(using data from previously published NTP bioassays) for the brominated trihalomethanes using
the Hill equation model. This analysis indicated that the shape of the dose response as well as the
Hill exponents were different for liver toxicity and tumorigenicity. The authors concluded that
these data do not support a causal relationship between liver toxicity/reparative hyperplasia and
tumor development.
Torti et al. (2001) conducted 1-week and 3-week inhalation exposure studies of
bromodichloromethane in wild type and transgenic mice. Bromodichloromethane toxicity was
transient. Regenerative lesions and increased labeling index were evident in the kidney cortex of
mice exposed to concentrations of 10 ppm and above for one week. After three weeks of
bromodichloromethane exposure, damaged areas of kidney cortex were entirely regenerated
(residual scarring was present) and labeling index measurements had returned to near baseline
levels. The study authors noted that these results are in contrast to those observed in similar
experiments performed with chloroform, where treatment of F344 rats and B6C3F, mice resulted
in continued cytotoxicity and elevated cell turnover for up to 90 days (Larson et al., 1996;
Templin et al., 1996). The mechanistic basis for these different responses to structurally similar
compound is unclear, but may reflect an induced change in metabolism or emergence of a resistant
cell population in animals treated with bromodichloromethane.
George et al. (2002) reported that exposure of male F344/N rats to bromodichloro-
methane in drinking water for two years at a level that significantly enhanced the prevalence and
multiplicity of hepatocellular adenomas and carcinomas had no effect on hepatocellular
proliferation. In the same study, the prevalence of renal tubular hyperplasia, but not tumor
incidence, was significantly increased at the high dose.
Based on an extensive evaluation of carcinogenicity data, cytotoxicity coupled with
regenerative hyperplasia is considered the primary mode of action for tumor formation following
exposure to high concentrations of chloroform, a structurally-related trihalomethane which has
low genotoxic potential (U.S. EPA, 2000d). However, two lines of evidence suggest that
chloroform is not a prototypical trihalomethane. First, the weight-of-evidence for at least two of
the brominated trihalomethanes indicates that they are genotoxic. This contrasts with the negative
weight of evidence evaluation for chloroform. Second, there is evidence that the brominated
trihalomethanes are readily bioactivated to mutagenic products via a glutathione ^'-transferase
mediated pathway, while chloroform is bioactivated only at very high concentrations. Therefore,
a common mode of action for carcinogenicity of chloroform and brominated trihalomethanes
cannot be assumed on the basis of current experimental evidence. Data to support a nonlinear
primary mode of action for tumor development in liver, kidney, and large intestine are currently
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lacking for the brominated trihalomethanes. In the absence of such information, combined with a
positive weight-of-evidence evaluation for genotoxicity, the mode of action for tumor
development is assumed to be a linear process.
D. Interactions and Susceptibilities
1. Potential Interactions
The toxicity of the brominated trihalomethanes appears to result from cytochrome P450-
mediated metabolism to reactive metabolites (U.S. EPA, 1994b). Therefore, agents which
increase or decrease the activity of enzymes responsible for metabolism of brominated
trihalomethanes may modify toxicity. Pankow et al. (1997) observed that pretreatment with
isoniazid or m-xylene (inducers of CYP2E1 and CYP2B1/CYP2B2, respectively) increased the
hepatotoxicity of dibromochloromethane in male rats, as measured by elevated serum leucine
aminopeptidase activity. Hewitt et al. (1983) observed that pretreatment with acetone, a CYP2E1
inducer, potentiated the acute toxicity of bromodichloromethane and dibromochloromethane in
male rats. Thornton-Manning et al. (1993) also found that pretreatment with acetone potentiated
the acute hepatotoxicity of bromodichloromethane in male rats. Conversely, the cytochrome
P450 inhibitor 1-aminobenzotriazole prevented bromodichloromethane-induced hepatotoxicity in
rats (Thornton-Manning et al. 1993). Current findings regarding the existence of glutathione-
mediated pathways for brominated trihalomethane metabolism (see sectionV.E.l) suggest that
treatments or agents which alter glutathione-»Y-transferase activity may potentially modify the
toxicity of brominated trihalomethanes.
The severity of brominated trihalomethane toxicity is potentially affected by the vehicle of
administration. Vehicle effects are well-documented in the toxicity of chloroform (e.g., Bull et al.
1986; Jorgenson et al. 1985) and there is some evidence that similar effects occur with brominated
trihalomethanes. In a study of vehicle effects on the acute toxicity of bromodichloromethane, a
high dose (400 mg/kg) of the chemical was more hepato- and nephrotoxic when given in corn oil
compared to aqueous administration, but this difference was not evident at a lower dose (200
mg/kg) (Lilly et al. 1994).
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2. Greater Childhood Susceptibility
A growing body of scientific evidence indicates that children may suffer disproportionately
from some environmental health risks. These risks may arise because the neurological,
immunological, and digestive systems of children are still developing (U.S. EPA, (1998a). In
addition, children may incur greater exposure because they eat more food, consume more fluids,
and breathe more air in proportion to their body weight when compared to adults (U.S. EPA,
1998a).
U.S. EPA (1998a) recently identified three key questions to consider when evaluating
health risks in children from exposure to drinking water disinfection byproducts (DBP) such as
the brominated trihalomethanes:
Is there information which shows that the disinfectant or DBP 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?
If the disinfectant or DBP causes cancer, are children more likely to be affected by
it than are adults?
If the disinfectant or DBP causes some noncancer toxic effect, are children more
likely to be affected by it than are adults?
The data available for evaluation of these issues as they relate to brominated trihalomethanes are
addressed below.
a. Effects on the fetus and ability to conceive and bear children
General Results from Animal Studies
Studies on the reproductive and developmental effects of brominated trihalomethanes are
summarized in Section V.E. At the present time, the available data suggest that
dibromochloromethane or bromoform cause effects in the developing fetus only at doses which
also produce signs of maternal toxicity. Furthermore, there is no evidence that indicates that
either of these chemicals impairs ability to conceive and bear children. Studies of these chemicals
in animals indicate that reproductive and developmental effects would likely occur only at doses
higher than those observed to cause liver and renal effects.
Bromodichloromethane has the most extensive database for developmental and
reproductive effect among the brominated trihalomethanes. Study results for the reproductive
and developmental effects of bromodichloromethane are mixed. No reproductive or
developmental effects were observed at doses up to approximately 116 mg/kg-day in females or
68 mg/kg-day in males in studies conducted in Sprague-Dawley rats (NTP, 1998). Adverse
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reproductive or developmental effect were not observed in rabbits exposed to doses as high as 55
or 76 mg/kg-day in drinking water on gestation days 6 to 29 (CCC, 2000c,d). Increased
incidences of sternebral aberrations (Ruddick et. al., 1983) and decreased ossification sites in the
forelimb and hindlimb (CCC, 2000b) have been observed in Sprague-Dawley rats administered
bromodichloromethane in corn oil and drinking water, respectively, at doses which induced
maternal toxicity.
Reproductive effects of bromodichloromethane have been noted in rodent assays.
Klinefelter et al. (1995) observed effects on sperm motility in rats administered 39 mg/kg-day in
drinking water for 52 weeks, but these effects were not accompanied by histopathological
changes in male reproductive tissues. Narotsky et al. (1997) observed a significantly increased
incidence of full litter resorption (FLR) in F344 rats treated with 75 mg/kg-day
bromodichloromethane by aqueous gavage throughout the period of organogenesis. This effect
was described as an all-or-nothing phenomenon, in that the litter was either fully resorbed or
appeared normal at parturition. This pattern was interpreted by the study authors as evidence for a
maternally-mediated mechanism, rather than a direct effect of bromodichloromethane on the
developing embryo. Bielmeier et al. (2001) observed increased incidence of FLR in F344 rats
treated with 75 or 100 mg/kg-day BDCM by aqueous gavage on one or more days during the
interval from gestation day 6 to 10. This response was strain-specific and FLR was not observed
in Sprague-Dawley rats treated with up to 100 mg/kg-day on gestation days 6 to 10.
Data from Epidemiological Studies
There is evidence from epidemiological studies which suggests that exposure to
bromodichloromethane is possibly associated with reproductive effects. An epidemiological study
by Waller et al. (1998) found an association between consumption of trihalomethanes in drinking
water and increased risk of spontaneous abortion. Analysis of personal exposure to individual
trihalomethanes demonstrated that bromodichloromethane had the strongest association to with
spontaneous abortion. When all four individual trihalomethanes (including chloroform) were
simultaneously included as variables in a logistic regression model, high personal
bromodichloromethane exposure had an odds ratio of 3.0 for spontaneous abortion (95% C.I. 1.4,
6.6). King et al. (2000) examined the relationship between stillbirth of fetuses weighing more
than 500 g and chlorination byproducts in drinking water. Increased risk of stillbirth was
associated with total trihalomethane concentration, chloroform concentration, and
bromodichloromethane concentration. The strongest association was observed for
bromodichloromethane, where risk doubled in women exposed to concentrations of 20 |ig/L or
more when compared to women consuming water containing concentrations less than 5 |ig/L.
Both population-based studies had reasonable cohorts, recruitment procedures, and
screening protocols. The number of subjects was adequate, and the information regarding fetal
losses was credible. Therefore, the conclusion that increased fetal loss in humans occurs in
association with the exposure measurements that were made is quite convincing. A weakness of
both population based studies, however, is that the measurement of the trihalomethanes appeared
to be at the source or at intermediate sources rather than at the tap. Thus, the
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bromodichloromethane concentrations that were measured could be different than the actual
exposures (levels at the tap) and/or the bromodichloromethane could be a surrogate for the actual
toxicant in both population-based studies. Although these human studies raise significant concern,
the occurrence of multiple disinfection byproducts in drinking water prevents the conclusion that
bromodichloromethane was the specific causative agent for the observed effects.
Full Litter Resorption in F344 Rats
As indicated above, two oral exposure studies of bromodichloromethane have observed
increased incidences of full litter resorption (FLR) in the F344 strain of rat. Narotsky et al.
(1997) observed a significantly increased incidence of full litter resorption (FLR) in F344 rats
treated with 75 mg/kg-day bromodichloromethane by aqueous gavage throughout the period of
organogenesis. These studies are of particular interest because of epidemiological association
between bromodichloromethane ingestion and increased risk of spontaneous abortion.
In light of human epidemiological studies possibly implicating bromodichloromethane in
pregnancy loss, Bielmeier et al. (2001) conducted a series of experiments to further elucidate the
mode of action of this compound on full litter resorption in F344 rats. Their experiments
investigated the effects of bromodichloromethane on serum levels of progesterone (necessary for
the maintenance of pregnancy) and luteinizing hormone (LH), which helps maintain the corpus
luteum and luteal function (the secretion of progesterone). In the rat, secretion of progesterone
by the corpus luteum is required up to GD 17 for maintenance of pregnancy. Luteal function is
established and maintained in the rat through a series of mechanisms that require cervical
stimulation by coitus, a pituitary response to this cervical stimulation, and a secretion of hormones
by the products of conception. In the newly formed corpus luteum, progesterone secretion is
autonomous. However, continued function requires luteotropic hormones secreted by the
pituitary. In many species LH is considered to be the most important luteotropic factor.
However, in rats prolactin has been identified as the primary luteotropic factor.
The hormonal requirements of the corpus luteum change with the physiological stage of
the pregnancy. Early in the rat pregnancy, prolactin secreted by the pituitary is essential for luteal
function. Starting at about GD 6, LH also is required for maintenance of luteal function in
addition to prolactin. The corpora lutea are dependent on LH until GD 11, at which time luteal
maintenance becomes dependent on one or more placental prolactin-like hormones (prolactogens)
(Gibori et al., 1988).
Bielmeier et al. (2001) investigated the effects of bromodichloromethane in rats during
two physiologically different stages of pregnancy: GD 6 to 10, which encompasses the LH-
dependent stage of pregnancy and GD 11 to 15, which is an LH-independent stage. These
authors observed increased incidences of FLR in F344 rats treated with 75 or 100 mg/kg-day
BDCM on one or more days during the LH-dependent interval from gestation day 6 to 10, but not
in rats treated during the LH-independent period. Thus, the critical period for induction of FLR
was limited to the luteinizing hormone (LH)-dependent phase of pregnancy, suggesting that
bromodichloromethane may disrupt pregnancy via a LH-mediated mode of action. The treatment
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period also overlapped the period when maintenance of pregnancy is dependent on progesterone
produced by the ovary. The response to bromodichloromethane was strain-specific and FLR was
not observed in Sprague-Dawley rats treated with up to 100 mg/kg-day on GD 6 to 10.
Measurement of serum LH and progesterone levels indicated that FLR was accompanied by a
marked reduction in progesterone concentration without a corresponding drop in LH levels. The
failure of bromodichloromethane to exert adverse effects after the LH-dependent window, the
reduction in serum progesterone level, and the unchanged serum luteinizing hormone levels led
this group to conclude that the target of toxicity was the ovary and that the mode of action was a
reduced sensitivity of the corpus luteum to luteinizing hormone.
Bielmeier et al. (2001) have emphasized that there are significant differences between rats
and humans in the hormonal maintenance of pregnancy. However, these authors have noted that
rats and humans are similar in that either LH or human chorionic gonadotropin (hCG), which act
via the same receptor, are required during specific gestational periods to maintain pregnancy. In
the opinion of these authors, the findings in rats may be relevant to observations of adverse
pregnancy outcomes in human epidemiological studies if bromodichloromethane disrupts
pregnancy by diminishing luteal responsiveness to LH via an effect on the LH/hCG receptor,.
However, additional research is required to determine whether the findings in rats are potentially
relevant to humans.
As noted by Narotsky et al. (1997), it is important to recognize that
bromodichloromethane-induced FLR occurs in F344 rats at doses several orders of magnitude
greater than would be experienced by humans consuming highly contaminated water. For
example, a dose of 75 mg/kg-day is approximately 15,000-fold higher than human intake
assuming a body weight of 70 kg, a bromodichloromethane concentration of 180 |ig/L, and water
consumption of 2L/day. However, the possibility that humans are more sensitive to
bromodichloromethane than rats cannot be dismissed. At this time, data are insufficient to
evaluate this possibility.
b. Childhood Cancer and Noncancer Effects
Bioactivation to reactive metabolites is an apparent prerequisite for toxicity and
carcinogenicity of the brominated trihalomethanes. Therefore, an important issue in the
assessment of childhood risk of cancer and other adverse effects is whether the enzymes
responsible for metabolism are more active in fetuses, neonates, and or children than in adults.
This section evaluates the available data for developmental expression and/or activity of three key
metabolizing enzymes that are known or anticipated to bioactivate the brominated
trihalomethanes: CYP2E1, CYP2B1/2 (in rodents only), and glutathione-»Y-transferase theta.
CYP2E1
Carcinogenicity of brominated trihalomethanes has been shown to be at least partly related
to bioactivation by the cytochrome P450 isoform CYP2E1 (U.S. EPA, 1994b). Thus, a higher
level of CYP2E1 activity in children relative to adults might predispose children to greater
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toxicity. Studies of human fetal liver have produced contradictory results, but suggest that
CYP2E1 protein is either not expressed or is expressed at levels lower than in adults (Hakkola et
al., 1998). Carpenter et al. (1996) detected immunoreactive CYP2E1 protein in liver samples
from fetuses ranging from 16 to 24 weeks in gestational age. The samples obtained were from
fetuses whose mothers did not have a history of alcohol use. The immunoreactive protein
exhibited a slightly lower molecular weight than observed for CYP2E1 from adult liver samples.
Expression of the corresponding mRNA was confirmed in a fetal liver sample of 19 weeks
gestational age by reverse transcriptase-polymerase chain reaction (RT-PCR). However,
CYP2E1 mRNA was not detectable in a fetal liver sample of 10 weeks gestational age, suggesting
(in the opinion of the study authors) that CYP2E1 expression may be related to specific stages of
fetal development. The catalytic capability of CYP2E1 protein in human fetal microsomes was
demonstrated by measuring the rate of ethanol oxidation to acetaldehyde. The rate of conversion
varied from 12 to 27% of that measured in adult microsomes. Treatment of fetal hepatocytes in
primary culture with ethanol or clofibrate indicated that fetal CYP2E1 protein is inducible
(approximately two-fold compared to untreated cells).
Viera et al. (1996) detected small amounts of CYP2E1 mRNA in fetal liver samples
(approximately 5 to 10% of the levels in adult liver) collected from fetuses aged 14 to 40 weeks.
However, these authors could not detect immunoreactive CYP2E1 protein in any of 27 fetal liver
samples Other studies have failed to detect either CYP2E1 protein or mRNA in fetal liver
samples. Cresteil et al. (1985) and Komori et al. (1989) did not detect immunoreactive protein or
mRNA in fetal liver samples of less than 16 weeks gestational age. Jones et al. (1992) did not
detect CYP2E1 mRNA or protein in liver samples that were of similar gestational age (16 to 18
weeks) to the samples examined by Carpenter et al. (1996). Juchau and Yang (1996) did not
detect CYP2E1 mRNA by RT-PCR in human embryonic tissues between days 45 and 60 of
gestation. The factors contributing to the different results are unknown, but may include inter-
individual variability, gestational age of the tissue examined (for the samples less than 16 weeks
gestational age), or the existence of factors other than developmental stage that control
expression.
Information on the presence of CYP2E1 in human fetal tissues other than the liver is
limited. Viera et al. (1998) examined the mRNA content of human fetal lung and kidney.
CYP2E1 mRNA was expressed at a very low level in both tissues and the levels remained stable
after birth. Studies of human fetal brain tissue indicate that CYP2E1 is expressed in human
embryonic brain tissue (see Juchau et al., 1994) and that relatively low levels of CYP2E1 mRNA,
immunoreactive protein, and catalytically active protein are present during the early fetal period of
development (Brzezinski et al., 1999). In one study, a dramatic increase in CYP2E1 was
observed at approximately gestation day 50, and a fairly constant level was maintained until at
least day 113 (Brzezinski et al., 1999). The relevance of the data for lung and brain is uncertain,
since these organs are not known to be targets for brominated trihalomethane toxicity.
Viera et al. (1996) investigated age-related variations in human CYP2E1 protein levels
and catalytic activity from birth through adulthood. These authors observed a rapid increase in
the immunoreactive CYP2E1 microsomal content within 24 hours after birth that was independent
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of the gestational age of the newborn. This activation was accompanied by a demethylation of
cytosine residues in the 5'-regulatory region of the gene, suggesting that methylation of specific
residues prevents transcription in the fetal liver. The CYP2E1 protein level gradually increased
during the first year and reached the adult level in children aged 1 to 10 years. CYP2E1 catalytic
activity was assessed by determination of in vitro hydroxylation of chlorzoxazone in 89
microsomal preparations. Chlorzoxazone hydroxylation activity increased within 24 hours after
birth and steadily increased during the first year. Catalytic activity reached adult levels at age 1 to
10 years.
Animal studies of CYP2E1 expression during development have given variable results.
One study indicated that CYP2E1 is expressed in the fetal rat liver and placenta and that levels are
increased in rat pups exposed to ethanol in utero or via lactation (Carpenter et al., 1997). Liver
samples from rat fetuses exposed to ethanol in utero showed a 2.4-fold increase in protein levels
and 1.5-fold increase in catalytic activity (Carpenter et al., 1997). Other authors have reported
that hepatic CYP2E1 gene transcription in rats is activated at birth and that the amount of
CYP2E1 reaches a peak prior to weaning (see Ronis et al., 1996). The protein level then falls to
approximately 25% of the peak level and remains stable into adulthood (Ronis et al., 1996).
The regulation of CYP2E1 is complex when examined at both the molecular (Lieber
1997) and physiological (Ronis et al. 1996) levels. The factors and processes responsible for the
increase in CYP2E1 protein levels and activity at birth have not been clearly identified. At the
physiological level, there is some evidence from rodent studies to suggest that growth hormone
regulates the constitutive expression of CYP2E1 (Ronis et al., 1996). The reduction of CYP2E1
from peak levels before weaning is reported to coincide with the increased levels of growth
hormone and with development of adult levels of growth hormone receptors (Ronis et al., 1996).
The occurrence of peak expression after birth has been attributed to a role of CYP2E1 in
gluconeogenesis, since there is a very high demand for energy production from glucose at this
developmental stage (see Ronis et al. 1996; Viera et al. 1996).
CYP2B1/2 (Rodents^)
Research conducted by Pankow et al. (1997) suggests that the closely-related CYP
isoforms 2B1 and 2B2 participate in the catabolism of dibromochloromethane in rats. These
isoforms show greater than 97% homology of amino acid sequence and have highly similar
genomic organization. To date, these isoforms have not been reported in adult or fetal human
tissues (Nelson et al., 1996; Juchau et al., 1998). Omiecinski et al. (1990) detected low levels of
CYP2B isoform mRNA in fetal rat liver on gestation day 15 (the earliest day in development
when the authors were able to macroscopically recognize and dissect the fetal liver) using the
polymerase chain reaction (PCR). Although the levels of mRNA expression were "substantially
lower" lower at day 15 than observed later in development, expression was clearly inducible by
pretreatment of pregnant rats with phenobarbital. Both constitutive and phenobarbital-induced
levels of mRNA increased with developmental age, reaching maximal levels at approximately
three weeks postpartum. No measurements of CYP2B activity were made in this study, so it is
not known whether changes in mRNA levels were paralleled by changes in catalytic activity.
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Juchau et al. (1998) reviewed a series of experiments that employed the selective substrate
probe pentoxyresorufin to test for CYP2B1/2 catalytic activity in fetal rat tissues. The overall
conclusion upon examination of all results was that if CYP2B isoforms are expressed in fetal rats,
they occur at biologically insignificant levels. Asoh et al. (1999) examined the catalytic activity of
CYP2B isoforms in fetal rat liver and found very low activity, a finding consistent with the
conclusion of Juchau et al. (1998).
Gebremichael et al. (1995) investigated the postnatal developmental profile of CYP2B1 in
Sprague-Dawley rats. CYP2B1 activity was detectable as early as seven days postnatally and
exhibited a variable pattern of expression (no clear trend evident) when assayed at Days 14, 21,
50, and 100. Asoh et al. (1999) examined the induction of CYP2B isoforms in neonatal rats. The
level of CYP2B catalytic activity was markedly higher at five days after birth relative to levels
observed in fetal hepatic tissue. Oral or intraperitoneal administration of phenobarbital to
pregnant rats increased the level of CYP2B expression and activity in neonates. Overall, these
findings suggest that CYP2B isoform activity is likely to be lower in fetuses than in neonates or
adults and that increased levels of activity may be observed in fetuses and neonates exposed to
inducing xenobiotics. The significance of this information for risk of cancer in human fetuses,
neonates, and children is uncertain since, as noted above, the CYP2B1/2 isoforms have not been
identified in humans.
Glutathione S-Transferase Theta
Genotoxicity studies in genetically engineered bacteria indicate that brominated
trihalomethanes can also be activated to mutagens by the product of the glutathione ^'-transferase
(GST) theta gene GSTT1-1 (DeMarini et al.,1997; Landi et al., 1999). Children and the fetus
could potentially experience increased risk of adverse effects if the activity of this enzyme was
higher at these life stages than in adults. Information on the developmental expression of GST
genes is currently limited. Although other classes of GSTs (alpha, mu, and pi) are expressed in
fetal liver, Mera et al. (1994) reported that theta-class GSTs were expressed in only adult liver.
This finding suggests that the fetus does not experience increased risk as a result of GST theta-
mediated mutagenicity. The occurrence of increased risk in children cannot be evaluated, since at
present the age at which synthesis of the GST theta is induced is unknown.
c. Childhood Cancers: Other Considerations
Examination of childhood cancer data compiled by the National Cancer Institute (Ries et
al. 1999) indicates that the incidence of hepatic, renal, and intestinal cancer (the types of cancer
observed in animal studies of brominated trihalomethane cancer potential) from causes other than
genetic predisposition are low. Primary neoplasms of the liver are rare in children younger than
15 years of age. There has been little change in the liver cancer incidence (total hepatic tumors) in
this age group over the last 21 years, with rates between 1.4 and 1.7 cases per million throughout
the time period. The incidence of hepatocellular carcinoma (the type of neoplasm observed in
mice treated with bromodichloromethane or dibromochloromethane) decreased in children
younger than 15 years of age during the period 1975 to 1995. The incidence rate of renal
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carcinoma remained very low (well under one case per million) in children younger than 15 years
during the period 1975 to 1995. Trend data were not available in Ries et al. (1999) for intestinal
cancer. These observations are consistent with (but do not prove) a conclusion that children are
not more susceptible than adults.
d. Conclusion
The available evidence for developmental expression of enzymes known to metabolize
brominated trihalomethanes supports the conclusion that children do not experience greater risk
from exposure to these compounds than do adults. At present there are no cancer incidence data
from humans to suggest that brominated trihalomethanes contribute to increased risk of cancer in
children.
3. Other Potentially Susceptible Populations
a. Subpopulations with altered levels of CYP2E1
CYP2E1 catalyzes the metabolism of brominated trihalomethanes to reactive intermediates
that mediate toxicity. Individuals with higher levels of CYP2E1 activity may therefore be at
greater risk for adverse health effects. This section describes factors associated with increased
levels of CYP2E1 activity and subpopulations who may be at increased risk as a result of these
factors.
Genetic Polymorphisms
Significant inter-ethnic differences exist in CYP2E1 polymorphism (Ronis et al., 1996;
Lieber 1997)) and it is possible that these differences could influence susceptibility to toxic
effects. The CYP2E1 polymorphisms currently reported in the literature are located in the 5'-
flanking (noncoding) regions of the gene, while the coding regions of the gene which specify
sequence appear to be well conserved among various ethnic groups (Ronis et al. 1996).
Mutations in the 5'-region of a gene can affect the regulation gene expression. The rare mutant c2
polymorphism of CYP2E1 is reported to be associated with higher transcriptional activity, protein
levels, and catalytic activity than the more common wild type allele (Lieber et al. 1997). As
reported by Lieber et al. (1997), the highest frequency of the c2 allele occurs in the Taiwanese
(0.28) and Japanese (0.19 to 0.27) populations. The frequencies in African-Americans,
European-Americans, and Scandinavians are much lower, generally in the range 0.01 and 0.05.
Efforts to link the occurrence of the c2 allele to higher rates of CYP2E1-mediated liver disease
have yielded inconsistent results. Thus, the functional significance of CYP2E1 polymorphism is
presently uncertain, and no conclusion can as yet be drawn about the relative risk for different
ethnic populations exposed to brominated trihalomethanes.
Altered Physiological or Health States
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The physiological functions of CYP2E1 include lipid metabolism and ketone utilization
(Lieber, 1997). Induction of CYP2E1 is observed in many conditions that elevate circulating
levels of lipids, including consumption of a high-fat or low-carbohydrate diet, starvation, obesity,
and insulin-dependent diabetes. Among the individuals likely to be affected by such conditions,
diabetics constitute the most clearly-defined susceptible population. Induction of CYP2E1 in
uncontrolled insulin-dependent diabetes is well-studied. In animals, this induction results in
elevated levels of CYP2E1 in the liver, kidney, and lung (Ioannides et al., 1996). Acetone (a
substrate of CYP2E1) is thought to be the inducing compound (Ronis et al., 1996). As a result of
induction, diabetic animals are more susceptible to the toxicity of some chemicals metabolized by
CYP2E1. While there are no specific data for the brominated trihalomethanes, this phenomenon
has been demonstrated for other halogenated compounds including chloroform, carbon
tetrachloride, trichloroethylene, and bromobenzene (Ioannides et al., 1996). Because the animal
and human orthologues of CYP2E1 show similar substrate specificity and bioactivation potential,
it is possible that diabetic humans may also be more susceptible to CYP2E1-mediated toxicity.
As CYP2E1 levels are reduced by insulin treatment, increased toxicity would be anticipated only
in poorly controlled or uncontrolled diabetics (Ioannides et al., 1996).
Alcohol consumption
CYP2E1 contributes to the metabolism of ethanol in humans and animals. Consumption
of ethanol induces CYP2E1 and chronic alcohol consumption is reported to result in as much as a
10-fold induction (Lieber, 1997). Hence, concurrent exposure to ethanol and brominated
trihalomethanes may increase susceptibility to adverse health effects. This interaction is of
concern because concurrent exposure to brominated trihalomethanes and ethanol is likely to occur
in a significant number of people. At present, there are no human or animal studies which
examine this interaction for brominated trihalomethanes. However, Wang et al. (1994) reported
that a single 100 mg/kg oral dose of ethanol administered to rats significantly increased the
toxicity of the structurally-related trihalomethane chloroform (also metabolized by CYP2E1).
Lieber (1997) noted that the hepatotoxicity of commonly used industrial solvents (e.g. carbon
tetrachloride, bromobenzene, and vinylidene chloride) and anesthetics (enflurane and halothane)
was increased in heavy drinkers, with a pattern of damage that was consistent with the selective
expression and induction of CYP2E1 in certain regions of the liver.
Concurrent exposure to other CYP2E1 inducers including pharmaceuticals
Because CYP2E1 is highly inducible by a wide range of xenobiotic compounds, prior
exposure to such inducers may potentially play a significant role in brominated trihalomethane
toxicity. Known inducers of CYP2E1 include certain therapeutic agents (acetaminophen,
isoniazid), volatile anaesthetics (halothane, isoflurane), and solvents (acetone, benzene, carbon
tetrachloride, trichloroethylene) (Raucy, 1995). Individuals exposed to or consuming these
inducers on a regular basis may therefore be at greater risk for brominated trihalomethane
toxicity.
b. Subpopulations with altered levels of glutathione ^'-transferase theta
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Individuals with Genetic Polymorphisms
Genotoxicity studies in bacteria (discussed in section V.F. 1) indicate that brominated
trihalomethanes can be activated to mutagens by the product of the glutathione ^'-transferase theta
gene GSTT1-1 (DeMarini et al.,1997; Landi et al., 1999). If similar pathways for bioactivation
exist in humans, GSTT1-1 polymorphism may influence susceptibility to brominated
trihalomethane-mediated toxicity. GSTT1-1 is characterized by a deletion polymorphism which
results in total loss of glutathione ^'-transferase-© activity in individuals (10 to 60% of the
population depending upon ethnicity and race) homozygous for the null genotype (GSTT1-1" ").
Individuals who are heterozygous for GSTT-1 (GSTTl-l+l~) have intermediate levels of enzyme
activity, while individuals homozygous for GSTT-1 (GSTTl-l+l+) have the highest levels. Landi
et al. (1999) have suggested that GSTT1-1+/+ individuals may experience excess genotoxic risk
when exposed to brominated trihalomethanes, particularly in organs which express glutathione-»Y-
transferase-theta and come in direct contact with brominated trihalomethanes. Potential target
sites would include the gastrointestinal tract and the bladder.
Concurrent Exposure to Inducers
If GSTT-1-mediated pathways for bioactivation of brominated trihalomethanes exist in
humans, factors which induce this enzyme may increase the risk of adverse health effects from
exposure. Although GSTT-1 is constitutively expressed, the level of its expression can be altered
by exposure to exogenous chemicals. Landi (2000) has summarized information on factors which
increase expression of the enzyme. In rats, aspirin increased GSTT-1 levels in the colon. Alpha-
tocopherol, coumarin; and other anticarcinogenic drugs increased gastric and esophageal levels;
and indole-3-carbinol and coumarin increased GSTT-1 levels in the liver. In mice, phenobarbital
induced hepatic GSTT-1 levels. Data for humans are limited, but there are indications that the
dietary intake of cruciferous vegetables enhances the expression of GSTT-1. It is possible that
consumption of these substances by GSTT-1 positive individuals could result in increased risk of
adverse effects. However, there are presently no data available for evaluation of this hypothesis.
c. Subpopulations with altered levels of putative protective compounds
Glutathione depletion has been observed to increase the hepatotoxicity of
bromodichloromethane (Gao et al., 1996). On the basis of these data, Gao et al. (1996) proposed
that populations with low baseline levels of glutathione (e.g., due to dietary deficiencies of
glutathione precursors such as cysteine and selenium) may be more sensitive to
bromodichloromethane-induced toxicity.
d. Possible gender differences
Apparent gender-related differences in the toxicity of brominated trihalomethanes have
been noted in studies where male and female animals were exposed concurrently (e.g. Aida,
1992a; Daniel, 1990; NTP,1987, 1989a; Tobe et al., 1982a). In general, male rats and mice
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appear to be somewhat more sensitive to the hepatic and renal toxicity induced by brominated
trihalomethanes than are females, although there are exceptions to this pattern (eg. the chronic
oral exposure study of bromoform conducted by NTP, 1989a in mice and the short-term study of
bromoform conducted by Aida et al., 1992a). While the basis for the apparent greater sensitivity
of males is unknown, the difference may be related to gender-specific differences in the level of
enzymes responsible for bioactivation of brominated trihalomethanes to toxic metabolites, or to
gender-specific differences in cellular protective mechanisms. It is important to note that at
present there is no evidence for gender-related differences in the activity levels of CYP2E1 or
GSTT-1 in humans.
E. Summary
It is generally believed that the toxicity of the brominated trihalomethanes is related to
their metabolism. This conclusion is based largely on the observation that liver and kidney, the
chief target tissues for these compounds, are also the primary sites of their metabolism. In
addition, treatments which increase or decrease metabolism also tend to increase or decrease
trihalomethane-induced toxicity in parallel.
Metabolism of brominated trihalomethanes is believed to occur via oxidative and reductive
pathways. Limited structure-activity data for brominated trihalomethanes and the structurally-
related trihalomethane chloroform suggest that bromination may influence the proportion of
compound metabolized via the oxidative and reductive pathways, with brominated compounds
being more extensively metabolized by the reductive pathway. Additional evidence suggests that
a GSH-mediated pathway may play an important role in metabolism of brominated
trihalomethanes. These data raise the possibility that brominated trihalomethanes may induce
adverse effects (toxicity and carcinogenicity) via several different pathways.
The precise biochemical mechanisms which link brominated trihalomethane metabolism to
toxicity have not been characterized, but many researchers have proposed that toxicity results
from the production of reactive intermediates. Reactive intermediates may arise from either the
oxidative (dihalocarbonyls) or the reductive (free radicals) pathways of metabolism. Such
reactive intermediates are known to form covalent adducts with various cellular molecules, and
may impair the function of those molecules and cause cell injury. Free radical production may
also lead to cell injury by inducing lipid peroxidation in cellular membranes. Direct evidence
showing a relationship between the level of covalent binding intermediates generated by either
pathway and the extent of toxicity is not available for the brominated trihalomethanes.
Manipulation of cellular glutathione levels suggests that this compound may play a protective role
in brominated trihalomethane-induced toxicity.
Individual brominated trihalomethanes have been shown to induce tumors in laboratory
animals. The mechanism by which brominated trihalomethanes induce tumors in target tissues has
not been fully characterized. DNA adducts can be formed by interaction of reactive metabolites
(resulting from oxidative and reductive metabolism) with DNA. In addition, preliminary evidence
suggested that DNA adducts can be formed through conjugation with glutathione and
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bioactivation of the resulting conjugates. Comparison of dose-response data for liver toxicity
(including cell proliferation) and tumorigenicity in mice suggests that tumor formation occurs at
concentrations lower than those which stimulate cell proliferation. No evidence for increased cell
proliferation in kidney was obtained in studies using doses up to 246 mg/kg-day for
bromodichloromethane, 312 mg/kg-day for dibromochloromethane, or 379 mg/kg-day for
bromoform.
Interaction with agents which increase or decrease the activity of enzymes responsible for
metabolism of brominated trihalomethanes may modify carcinogenicity/toxicity. Pretreatment
with inducers of CYP2E1 has been observed to increase the hepatotoxicity of
bromodichloromethane and dibromochloromethane in male rats. Pretreatment with m-xylene, an
inducer of the CYP2B1/CYP2B2 isoforms, increased the hepatotoxicity of
dibromochloromethane in male rats. Conversely, administration of the cytochrome P450 inhibitor
1-aminobenzotriazole prevented bromodichloromethane-induced hepatotoxicity in rats. Recent
findings indicating possible glutathione-mediated metabolism of brominated trihalomethanes
suggest that treatments or agents which alter glutathione-»Y-transferase activity could potentially
modify the toxicity of brominated trihalomethanes.
The severity of brominated trihalomethane toxicity is potentially affected by the vehicle of
administration. In a study of vehicle effects on the acute toxicity of bromodichloromethane, a
high dose (400 mg/kg) of the chemical was more hepato- and nephrotoxic when given in corn oil
compared to aqueous administration, but this difference was not evident at a lower dose (200
mg/kg).
A number of potentially sensitive subpopulations have been identified for health effects of
brominated trihalomethanes. A growing body of scientific evidence indicates that children may
suffer disproportionately from some environmental health risks. These risks may arise because the
neurological, immunological, and digestive systems of children are still developing. In addition,
children may incur greater exposure because they eat more food, consume more fluids, and
breathe more air in proportion to their body weight when compared to adults. U.S. EPA has
identified three key questions to consider when evaluating health risks to children from drinking
water disinfection byproducts (DBP), including the brominated trihalomethanes: 1) Is there
information which shows that the DBP causes effects in the developing fetus or impairs ability to
conceive and bear children? 2) If the DBP causes cancer, are children more likely to be affected
by it than are adults? and 3) If the DBP causes a noncancer toxic effect, are children more likely
to be affected by it than are adults? The available data for dibromochloromethane and bromoform
suggest that developmental effects in animals occur only at doses which cause maternal toxicity
and at doses lower than those which induce histopathological effects in the liver and kidney.
There is no evidence that these compounds impair the ability to conceive or have children. In
animal studies, exposure to bromodichloromethane resulted in reduced sperm motility; this effect
was not accompanied by histopathologic changes in the male reproductive system. Exposure of
pregnant F344 rats during gestation days 6-9 caused full litter resorption. This response was not
observed in similarly exposed pregnant Sprague-Dawley rats. Epidemiological studies have found
an association between exposure to bromodichloromethane in drinking water and increased
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spontaneous abortion and increased stillbirth. Although these studies raise concern for human
health effects, the occurrence of multiple disinfection byproducts in drinking water prevents the
conclusion that bromodichloromethane is the causative agent.
At present, there are no cancer data which indicate that brominated trihalomethanes
contribute to increased risk of cancer in children. No studies were located which examined pre-
or post-pubertal cancer rates in humans in relation to brominated trihalomethane exposure.
Cancer bioassays of brominated trihalomethanes conducted in mice and rats have not used study
designs that included perinatal exposure.
The available evidence suggests that the toxic effects of brominated trihalomethanes are
mediated by the enzymes CYP2E1, CYP2B1/2 (in rodents), and glutathione S-transferase theta
(GSTT-1). The weight of evidence from studies of the developmental expression of these
enzymes supports the conclusion that children do not experience greater risk from brominated
trihalomethane exposure as a result of higher metabolic activity.
In addition to children, other potentially sensitive populations include those with altered
levels or activity of CYP2E1 or GSTT-1 and those with altered levels of glutathione. Factors
contributing to increases in CYP2E1 activity potentially include genetic polymorphisms; altered
physiological or health states; alcohol consumption; and concurrent exposure to other inducers,
including some pharmaceuticals and solvents. Factors contributing to increased GSTT-1 activity
include genetic polymorphisms and concurrent exposure to inducers. Based on observations in
animals, human populations with reduced levels of glutathione as a result of dietary deficiency or
other factors may experience increased sensitivity to the toxic effects of bromodichloromethane.
Apparent gender-related differences in the toxicity of brominated trihalomethanes have
been noted in studies where male and female animals were exposed concurrently. In general, male
rats and mice appeared to be more sensitive than females to liver and renal toxicity, although
some exceptions to this pattern have been noted. There is no evidence for a similar pattern of
gender response in humans.
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VIII. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
This section quantifies the toxicological effects of brominated trihalomethanes based on
health effects information presented in Sections V and VI. At present, there are two basic
approaches to quantification of toxicological effects: the conventional NOAEL/LOAEL approach
and benchmark dose modeling. Benchmark dose (BMD) modeling (U.S. EPA, 1995; 2000b) was
chosen as the preferred approach for quantifying toxicological effects of the brominated
trihalomethanes. BMD modeling avoids several limitations of the NOAEL/LOAEL approach,
including: 1) the slope of the dose-response plays little role in determining the NOAEL; 2) the
NOAEL (or LOAEL) is limited to the doses tested experimentally; 3) the determination of the
NOAEL is based on scientific judgement, and is subject to inconsistency; and 4) experiments
using fewer animals tend to produce larger NOAELs, and as a result may produce larger health
advisories (HAs) or reference doses (RfDs) (U.S. EPA, 1995) that may not be sufficiently
protective of human health. In contrast, benchmark doses (BMDs) are not limited to the
experimental doses, appropriately reflect the sample size, and can be defined in a statistically
consistent manner. The BMD approach was therefore selected for quantification of toxicological
effects of the brominated trihalomethanes. Values for HAs and RfDs derived using the
conventional NOAEL/LOAEL approach are presented in the text for comparison with those
obtained using the BMD approach.
The methods employed for BMD modeling are described in Appendix A. The modeling
was performed using the BMDS software (Version 1.2) developed by the U.S. EPA National
Center for Environmental Assessment. The BMDs and BMDLs were calculated based on a BMR
of 10% extra risk for all quantal endpoints analyzed. For continuous data, the BMR was defined
as 1.1 standard deviations, which corresponds to an additional risk of approximately 10% when
the background response rate is assumed to be 1% with normal variation around the mean and
constant standard deviation (Crump, 1995). The BMDL10 was defined as the 95% lower bound
on the corresponding BMD estimate. Confidence bounds were automatically calculated by the
BMDS software using a likelihood profile method.
A. Bromodichloromethane
1. Noncarcinogenic effects
a. One-day Health Advisory
Studies of the acute toxicity of bromodichloromethane are summarized in Table VIII-1.
Lilly et al. (1994) administered single doses of bromodichloromethane by either oil or aqueous
gavage to male F344 rats at dose levels of 200 or 400 mg/kg. This study identified a LOAEL of
200 mg/kg-day based on histologic lesions in the kidney and changes in urinary parameters. A
NOAEL value was not identified for either vehicle. Data for hepatic vacuolar degeneration and
renal tubular degeneration obtained using the aqueous vehicle were modeled using the BMDS
software. BMD and BMDL10 values of 263 and 182 mg/kg-day, respectively, were calculated
using the hepatic data. BMD and BMDL10 values of 131 and 8.9 mg/kg-day, respectively, were
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VIII- 1
February 20, 2003
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obtained using the renal data. The BMDL10 for renal tubular degeneration is the lowest calculated
across studies, but is not considered a reliable estimate because there is insufficient information to
accurately characterize the shape of the dose-response curve in the region of interest.
Thornton-Manning et al. (1994) administered bromodichloromethane to female F344 rats
by aqueous gavage for five consecutive days at dose levels ranging from 75 to 300 mg/kg-day.
This study identified a NOAEL of 75 mg/kg-day and a LOAEL of 150 mg/kg-day based on
increased liver and kidney weights and histologic lesions in the liver (mild centrilobular
hepatocellular vacuolar degeneration) and in the kidney (mild renal tubule vacuolar degeneration).
An analogous study (Thornton-Manning et al., 1994) conducted in female C57BL/6J mice
indicated that the mice were less sensitive to bromodichloromethane than the rats, as no
treatment-related histologic lesions were observed in the liver or kidney. However, similar
NOAEL and LOAEL values were identified based on increased liver weight and changes in serum
chemistry parameters. Data for renal tubular degeneration in rats were analyzed using the BMD
approach. BMD and BMDL10 values of 133 and 65 mg/kg-day, respectively, were calculated for
this endpoint. The BMD is in close agreement with the BMD value calculated for the same
endpoint using the data of Lilly et al. (1994).
Two reproductive studies which examined full litter resorption were also considered for
derivation of the One-day HA. Bielmeier et al. (2001) examined the occurrence of full litter
resorption in F344 rats treated with 0, 75 or 100 mg/kg-day bromodichloromethane by aqueous
gavage on gestation day 9. The LOAEL for this effect was 75 mg/kg-day. Narotsky et al. (1997)
evaluated the same endpoint in F344 rats administered 0, 25, 50, or 75 mg/kg-day on gestation
days 6 through 15. Full litter resorption was observed at 50 and 75 mg/kg-day. The NOAEL and
LOAEL in this study were thus identified as 25 and 50 mg/kg-day, respectively. When data from
these studies were analyzed using the BMD approach, BMD values of 48 and 23 mg/kg-day were
obtained for the Narotsky et al. (1997) and Bielmeier et al. (2001) studies, respectively. The
higher value from the Narotsky et al. (1997) study was considered the more reliable estimate of
the BMD because it was based on response data that included lower doses, one of which was an
apparent NOAEL. The BMDL10 calculated from the Narotsky et al. (1997) data was 30 mg/kg-
day.
Three additional studies were considered as candidates for derivation of the One-day HA.
Lilly et al. (1997) administered single doses of bromodichloromethane by aqueous gavage to male
F344 rats at dose levels ranging from 123 to 492 mg/kg. Based on changes in urinary parameters,
this study identified a NOAEL of 164 mg/kg-day and a LOAEL of 246 mg/kg-day. No
histopathological examination was conducted in this study. The study by French et al. (1999),
which investigated immune system response, identified a similar NOAEL value. However, the
database for immune response to bromodichloromethane is limited when compared to information
on hepatic and renal toxicity. Adverse effects were noted only at the highest dose and frank effect
level, and evidence for vehicle effects on immunotoxicity endpoints was observed. Keegan et al.
(1998) administered single doses of bromodichloromethane to male F344 rats by gavage at dose
levels ranging from 21 to 246 mg/kg. The study authors identified a NOAEL of 41.0 mg/kg-day
and a LOAEL of 81.9 mg/kg-day based on elevations in serum markers of hepatotoxicity (ALT,
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AST, and SDH). Histopathological examination was not conducted and this was considered to be
a limitation of the investigation. These three studies were not considered further for derivation of
the One-day Health Advisory (HA), and thus data reported in them were not analyzed using the
BMD approach.
The study conducted by Narotsky et al. (1997) was selected for derivation of the One-day
HA. The critical effect in this study was full litter resorption observed in pregnant F344 rats
treated with bromodichloromethane. The BMDL10 value calculated for this endpoint was 30
mg/kg-day, which is roughly half of the most reliable BMDL10 value calculated for
histopathological changes in kidney (Thornton-Manning et al., 1994). Although dosing in the
Narotsky et al. (1997) study lasted from gestation days 6 through 15, a subsequent study by
Bielmeier et al. (2001) indicated that a single dose (75 mg/kg) of bromodichloromethane on
gestation day 9 was sufficient to elicit full litter resorption in the same strain of rats. Since there is
presently insufficient information available to fully assess the occurrence of reproductive effects in
humans exposed to bromodichloromethane, use of data for full litter resorption was adopted as a
conservative approach to derivation of the One-day HA. The One-day HA for a 10-kg child is
calculated using the following equation:
(30 mg/kg-day) (10 kg)
One-day HA = = l.Omg/L
(300) (1 L/day)
where:
30 mg/kg-day = BMDL10 based on incidence of full litter resorption in F3 44
rats treated with bromodichloromethane on gestation days 6
to 15.
10 kg = Assumed body weight of a child
300 = Uncertainty factor based on NAS/OW guidelines. This
value includes a factor of 10 to protect sensitive human
populations; a factor of 10 for extrapolation from animals
to humans; and a factor of 3 to account for database
limitations and uncertainty regarding possible reproductive
effects of bromodichloro-methane in humans
1 L/day = Assumed water consumption of a 10-kg child
For comparative purposes, the One-day HA derived using the conventional
NOAEL/LOAEL approach would also be based on data from the Narotsky et al. (1997) study.
This study identified a NOAEL of 25 mg/kg-day based on FLR, which was the lowest value
among the candidate studies. Using this NOAEL and an uncertainty factor of 300 as described
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Table VIII-1 Summary of Candidate Studies for Derivation of the One-day HA for Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
mg/kg-day
LOAEL
mg/kg-day
BMD
mg/kg-day
BMDL10
mg/kg-day
Lilly et al.
(1994)
Rat
F344
M
6
0
200
400
Gavage
(oil)
Single Dose
Body, liver, and
kidney weights,
serum and urine
chemistry, liver and
kidney histology
200
(minimal renal
tubule
degeneration and
necrosis,
changes in
urinary
parameters)
Not modeled
Lilly et al.
(1994)
Rat
F344
M
6
0
200
400
Gavage
(aqueous)
Single Dose
Body, liver, and
kidney weights,
serum and urine
chemistry, liver and
kidney histology
"¦
200
(minimal renal
tubule
degeneration,
changes in
263
182
(Hepatic
vacuolar
degeneration in
males)
urinary
parameters)
131
8.9
(Renal tubule
degeneration in
males)
Lilly et al.
(1997)
Rat
F344
M
5
0
123
164
246
328
492
Gavage
(aqueous)
Single Dose
Body, liver, and
kidney weights,
serum and urine
chemistry
164
246
(changes in
urinary
parameters)
Not modeled
Keegan et al.
(1998)
Rat
F344
M
6
0
21
31
41
82
123
164
246
Gavage
(aqueous)
Single Dose
Body, liver, and
kidney weights,
serum chemistry
41
82
(elevated ALT,
AST, and SDH
activities)
Not modeled
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Reference
Species
n
Dose
Route
Exposure
Endpoints
NOAEL
LOAEL
BMD
BMDL10
Sex
Duration
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
French et al.
Rat
6
0
Gavage
5 days
Body, spleen, and
150
300 (FEL)
Not modeled
__
(1999)
C57BL/6
75
(aqueous)
thymus weights,
(mortality,
F
150
immune function
decreased body
300
weight, altered
immune
response)
Thornton-
Rat
6
0
Gavage
5 days
Body, liver, and
75
150
133
65
Manning et
F344
75
(aqueous)
kidney weights,
(increased liver
(renal tubular
al. (1994)
F
150
serum chemistry,
and kidney
degeneration)
300
liver and kidney
weights, mild
histology
centrilobular
hepatocellular
vacuolar
degeneration,
mild renal tubule
vacuolar
degeneration)
Thornton-
Mouse
6
0
Gavage
5 days
Body, liver, and
75
150
Not modeled
Manning et
C57BL/
75
(aqueous)
kidney weights,
(increased liver
al. (1994)
6J
100
serum chemistry,
weight, elevated
F
liver and kidney
ALT and SDH
histology
activities)
Narotsky et
Rat
12-
0
Gavage
Gestation
Body weight,
25
50
48
30
al.
F344
14
25
(oil)
days 6-15
clinical signs,
(full-litter
(full-litter
(1997)*
F
50
(water)**
developmental
resorption)
resorption)
75
parameters
Bielmeier et
Rat
8-
0
Gavage
Gestation day
Full litter resorption;
..
75
23
4.2
al. (2001)*
F344
11
75
(aq)
9
hormone profiles
(full-litter
(lull-litter
F
100
resorption)
resorption)
* The NOAEL and LOAEL values listed are for reproductive or developmental effects.
** The NOAEL and LOAEL values were the same in either vehicle. BMD modeling was performed on aqueous vehicle data only.
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above, the One-day HA for a 10-kg child calculated using the conventional approach would be
0.8 mg/L (rounded from 0.83 mg/L).
b. Ten-day Health Advisory
Sixteen studies were considered for derivation of the Ten-day HA for
bromodichloromethane. These studies are summarized in Table VIII-2 below. Aida et al.
(1992a) administered microencapsulated bromodichloromethane in the diet to Wistar rats for one
month at dose levels ranging from 20.6 to 203.8 mg/kg-day. This study identified a NOAEL of
61.7 mg/kg-day and a LOAEL of 189.0 mg/kg-day in male rats based on histologic changes in the
liver (swelling of hepatocytes, single cell necrosis, hepatic cord irregularity, and bile duct
proliferation). Analysis using the BMD approach calculated BMD and BMDL10 values of 34 and
17 mg/kg-day, respectively, based on data for liver cell vacuolization in females.
Data from four of the other candidate studies are consistent with the histopathological
results obtained by Aida et al. (1992a). Melnick et al. (1998) administered
bromodichloromethane by gavage to female B6C3F, mice for 5 days/week for 3 weeks and
identified a NOAEL of 75 mg/kg-day (duration-adjusted NOAEL of 54 mg/kg-day) and a
LOAEL of 150 mg/kg-day (duration-adjusted LOAEL of 107 mg/kg-day) based on histologic
changes in the liver (hepatocyte hydropic degeneration). Analysis using the BMD approach
calculated duration-adjusted BMD and BMDL10 values of 31 and 8.4 mg/kg-day, respectively, for
this endpoint. Condie et al. (1983) administered bromodichloromethane by gavage to male CD-I
mice for 14 days and identified a NOAEL of 74 mg/kg-day and a LOAEL of 148 mg/kg-day
based on minimal to moderate liver and kidney lesions. Analysis using the BMD approach
calculated BMD and BMDL10 values of 24 and 7.5 mg/kg-day, respectively, based on data for
histopathological changes in the liver. NTP (1998) conducted histopathologic examinations in
conjunction with a study of reproductive and developmental toxicity in Sprague-Dawley rats.
Although no reproductive or developmental toxicity was observed at the dose levels investigated,
histopathological changes were noted in the liver of males rats treated with the compound for 35
days. The NOAEL and LOAEL for this effect were 9 and 38 mg/kg-day, respectively. Analysis
of data for single cell hepatic necrosis using the BMD approach calculated BMD and BMDL10
values of 35 and 18 mg/kg-day, respectively, which are virtually identical to the values calculated
using the liver cell vacuolization data for females from the Aida et al. (1992b) study. Coffin et al.
(2000) observed hydropic degeneration in female mice treated with 150 mg/kg-day
bromodichloromethane in corn oil for 11 days. These data were not modeled because other
studies utilized doses lower than 150 mg/kg-day, which allowed better characterization of
response in the low dose region of the dose-response curve.
In contrast to the studies described above, Chu et al. (1982a) did not observe any
microscopic lesions in the liver when Sprague-Dawley rats were administered
bromodichloromethane at doses up to 68 mg/kg-day in the drinking water for 28 days. The
studies of Munson et al. (1982) and NTP (1987) did not conduct histopathological examinations.
These studies identified NOAELs ranging from 50 to 150 mg/kg-day and LOAELs ranging from
125 to 300 mg/kg-day for other endpoints, including depressed humoral immunity (Munson et al.,
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1982), decreased weight gain (NTP, 1987; rats), and increased mortality and gross renal
pathology (NTP, 1987; mice). These data were not analyzed by the BMD approach.
Seven studies that examined developmental and/or reproductive endpoints were evaluated.
Two studies reported the incidence of full litter resorption in F344 rats following treatment with
bromodichloromethane. Bielmeier et al. (2001) examined the occurrence of full litter resorption
in F344 rats treated with 0, 75 or 100 mg/kg-day bromodichloromethane by aqueous gavage on
gestation day 9. The LOAEL for this effect was 75 mg/kg-day. Narotsky et al. (1997) evaluated
the same endpoint in F344 rats administered 0, 25, 50, or 75 mg/kg-day on gestation days 6
through 15. Full litter resorption was observed at 50 and 75 mg/kg-day. The NOAEL and
LOAEL in this study were thus identified as 25 and 50 mg/kg-day, respectively. When data from
these studies were analyzed using the BMD approach, BMD values of 48 and 23 mg/kg-day were
obtained for the Narotsky et al. (1997) and Bielmeier et al. (2001) studies, respectively. The
higher value from the Narotsky et al. (1997) study was considered the more reliable estimate of
the BMD because it was based on response data that included lower doses, one of which was an
apparent NOAEL. The BMDL10 calculated from the Narotsky et al. (1997) data was 30 mg/kg-
day. Ruddick et al. (1983) observed an increased incidence of sternebral aberrations in the pups
of Sprague-Dawley rats administered bromodichloromethane in corn oil by gavage. Statistical
analysis of the published data indicated that the NOAEL and LOAEL for this effect were 100
mg/kg-day and 200 mg/kg-day, respectively. The BMD and BMDL10 obtained for this study
were 27 and 15 mg/kg-day, respectively.
The remaining reproductive/developmental studies sponsored by the Chlorine Chemistry
Council (CCC) were also evaluated. CCC (2000a, b) examined developmental toxicity in New
Zealand rabbits and identified developmental NOAELs of 76 and 55 mg/kg-day (the highest doses
tested in each study). The CCC (CCC, 2000c,d) also examined reproductive and developmental
toxicity in Sprague Dawley rats. In a range-finding study (CCC, 2000c), F, generation pups
exposed to bromodichloromethane via lactation and possibly by consumption of water supplied to
the dams exhibited reduced body weights and body weight gains. These effects occurred at
exposure levels which also resulted in maternal toxicity. Biologically meaningful average daily
doses could not be established in this experiment; therefore, the concentration-based NOAEL and
LOAEL values for developmental effects were 50 and 150 ppm. based on changes in F, pup body
weight and body weight gain. In a subsequent developmental study, CCC (2000d) identified
NOAEL and LOAEL values of 45 and 82 mg/kg-day, respectively, based on decreased number of
ossification sites per fetus for the forelimb phalanges and hindlimb metatarsals and phalanges.
This effect was observed at doses associated with maternal toxicity. Endpoints from these studies
were not modeled because other studies identified adverse effects at lower doses.
Data for maternal toxicity from three reproductive/developmental studies were also
considered for derivation of the 10-day HA for bromodichloromethane. Narotsky et al. (1997)
observed decreased maternal body weight gain on gestation days 6 to 8 in female rats
administered 25 mg/kg-day (the lowest dose tested) by aqueous gavage in 10% Emulphor. BMD
modeling identified BMD and BMDL10 values of 18 and 10 mg/kg-day, respectively, for this
endpoint. The data reported in this study did not permit evaluation of body weight or body
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weight gain at other time points during the treatment period. CCC (2000d) reported decreased
maternal body weight gain at several time points in pregnant rats administered
bromodichloromethane in drinking water, with the most severe effect observed immediately after
initiation of treatment on gestation days 6 to 7. The NOAEL and LOAEL for decreased body
weight on gestation days 6 to 7 were 18.4 and 45 mg/kg-day, respectively. When body weight
gain for this interval was modeled, the resulting BMD and BMDL10 values were approximately 18
and 15 mg/kg-day, respectively. However, the modeled fits to the data were poor, and the results
were not considered sufficiently reliable for derivation of a health advisory. To address this
problem, body weight gain data for gestation days 6 to 9 were also modeled. Reliable values of
23 and 18 mg/kg-day were obtained for the BMD and BMDL10, respectively. The CCC (2000b)
study observed decreased maternal body weight gain at several time points in pregnant rabbits
administered bromodichloromethane in the drinking water on gestation days 6 to 29. The
NOAEL and LOAEL for decreased maternal body weight gain (corrected for gravid uterine
weight) on gestation days 6 to 21 were 13.4 and 35.3 mg/kg-day, respectively. A BMD value of
50 mg/kg-day was obtained for this data set, but the BMDS software failed to identify the
corresponding BMDL10. No further modeling was attempted since this value was well above the
lowest BMDs obtained in some other candidate studies.
As evident from the data in Table VIII-2, the four studies that examined histopathological
changes in the liver are in close agreement, having identified BMD values ranging from 24 to 35
mg/kg-day. The corresponding BMDL10 values ranged from 7.5 to 18 mg/kg-day. Maternal
toxicity occurred in rats at similar levels in two developmental studies. These studies identified
BMD values of 18 and 23 mg/kg-day, with corresponding BMDL10 values of 10 and 18 mg/kg-
day. The NTP (1998) and CCC (2000d) drinking water studies were selected to derive the Ten-
day HA. Selection of these studies was based on the administration of bromodichloromethane in
drinking water, the most relevant route of exposure. In addition, these studies utilized a lower
range of doses, which provided information on the shape of the dose-response curve in the region
of interest. The Ten-day HA is calculated according to the following equation:
(18 mg/kg-day) (10 kg)
Ten-day HA = = 0.60 mg/L (rounded to 0.6 mg/L)
(300) (1 L/day)
where:
18 mg/kg-day = BMDL10 based on single cell hepatic necrosis in rats
administered bromodichloromethane in the drinking water
for 35 days or decreased maternal body weight gain on
gestation days 6-9 in pregnant female rats administered
bromodichloromethane in the drinking water.
10 kg = Assumed body weight of a child
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300 = Uncertainty factor based on NAS/OW guidelines. This
value includes a factor of 10 to protect sensitive human
populations and a factor of 10 for extrapolation from
animals to humans, and a factor of 3 to account for
database limitations and uncertainty regarding possible
reproductive effects of bromodichloromethane in humans
1 L/day = Assumed water consumption of a 10-kg child
For comparative purposes, the Ten-day HA derived using the conventional NOAEL/LOAEL
approach would be based on data from the CCC (2000d) study. This study identified a NOAEL
of 18 mg/kg-day based on reduced maternal body weight gain in pregnant rabbits. Using this
NOAEL and an uncertainty factor of 300 as described above, the Ten-day HA for a 10-kg child
calculated using the conventional approach would be 0.60 mg/L (rounded to 0.6 mg/L).
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Table VIII-2 Summary of Candidate Studies for Derivation of the Ten-day HA for Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL1C
(mg/kg-day)
Aida et al.
(1992a)
Rat
Wistar
M, F
7
Male
0
21
62
189
Females
0
21
66
204
Feed
1 month
Clinical signs, body
weight, serum
chemistry,
hematology, histology
62
189
(liver
histopathology
in males)
34
17
(liver cell
vacuolation in
females)
Chu et al.
(1982a)
Rat
SD
M
10
0
0.8
8
68
Drinking
water
28 days
Clinical signs, serum
chemistry, histology
68
--
No data to model
--
Condie et al.
(1983)
Mouse
CD-I
M
8-
16
0
37
74
148
Gavage
(oil)
14 days
Serum enzymes, PAH
uptake in vitro,
histology
74
148
(elevated ALT,
decreased PAH
uptake, liver and
kidney
histopathology)
24
7.5
(hepatic
centrilobular
pallor)
125
53
(Renal epithelial
hyperplasia)
Melnick et
al.
(1998)
Mouse
B6C3Fj
F
10
0
75
150
326
Gavage
(oil)
3 weeks
(5 d/wk)
Body and liver
weights, serum
chemistry, liver
histology
75
150
(liver
histopathology)
31 f
8.4
(Hepatocyte
hydropic
degeneration)
Munson et
al.
(1982)
Mouse
CD-I
M, F
8-
12
Males
0
50
125
250
Gavage
(aq.)
14 days
Body and organ
weights, serum
chemistry,
hematology, and
immune function
50
125
(depressed
humoral
immunity)
Not modeled
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VIII - 10 February 20, 2003
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Table VIII-2 (cont.)
Reference
Species
n
Dose
Route
Exposure
Endpoints
NOAEL
LOAEL
BMD
BMDL10
Sex
Duration
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
NTP
Rat
5
0
Gavage
14 days
Body weight, clinical
150
300
Not modeled
__
(1987)
F344/N
38
(oil)
signs, gross necropsy
(decreased
M, F
75
weight gain)
150
300
600
NTP
Mouse
5
0
Gavage
14 days
Body weight, clinical
75
150 (FEL)*
Not modeled
__
(1987)
B6C3Fj
19
(oil)
signs, gross necropsy
(mortality,
M, F
38
lethargy, gross
75
renal pathology)
150
300
Coffin et al.
Mouse
10
0
Gavage
11 days
Relative liver wt.,
150
Not modeled
(2000)
B6C3Fj
150
(oil)
liver histopathology;
F
300
labeling index
NTP
Rat
5-
0
Drinking
35 days
Body and organ
9
38
35
18
(1998)
SD
13
9
water
weights, serum
(liver
(liver cell
M
38
chemistry,
histopathology)
necrosis)
(group A)
67
hematology, gross
necropsy, histology,
sperm evaluation
Ruddick et
Rat
9-
0
Gavage
Gestation
Body and organ
100
200
27
15
al. (1983)*
SD
14
50
(oil)
days 6-15
weights; maternal
(sternebral
(sternebral
F
100
serum chemistry;
aberrations)
aberrations)
200
hematology, and
histopathology;
developmental
parameters
Draft - Do Not Cite or Quote
VIII - 11 February 20, 2003
-------
Table VIII-2 (cont.)
Reference
Species
n
Dose
Route
Exposure
Endpoints
NOAEL
LOAEL
BMD
BMDL10
Sex
Duration
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
Narotsky et
Rat
12-
0
Gavage
Gestation
Body weight, clinical
25
50
48
30
al. (1997)*
F344
14
25
(oil)
days 6-15
signs, developmental
(developmental)
(full-litter
(full-litter
F
50
(Emul-
parameters
resorption)
resorption)
75
phor)**
-
25
18
10
(maternal)
(reduced
(reduced
matermal body
matermal body
weight gain
weight gain
gestation days
gestation days 6-
6-8, aqueous
8, aqueous
vehicle only)
vehicle only)
Bielmeier et
Rat
8-
0
Gavage
Gestation
Full litter resorption;
..
75
23
4.2
al. (2001)*
F344
11
75
(aq)
day 9
hormone profiles
(full-litter
(full-litter
F
100
resorption)
resorption)
ccc
Rat
10
0 ppm
Drinking
Males 64
Reproductive and
50 ppm
150 ppm
Not modeled
__
(2000c)*
SD
50 ppm
water
days
developmental
(reduced ¥1 pup
M, F
150 ppm
parameters
weight and
450 ppm
Females 74
weight gain)
1350 ppm
days
CCC
Rat
25
0.0
Drinking
Gestation
Reproductive and
45.3
82.0
Not modeled
..
(2000d)*
SD
2.2
water
days 6-21
developmental
(developmental)
(reduced
F
18.4
parameters
number of
45.0
ossification sites
82.0
in phalanges or
metatarsals
occurring with
maternal
toxicity)
18.4
45
23
18
(maternal)
(reduced
(reduced
maternal body
maternal body
weight gain
weight gain
gestation days
gestation days 6-
6-7)
9; see text for
comment)
Draft - Do Not Cite or Quote
VIII - 12 February 20, 2003
-------
Table VIII-2 (cont.)
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day)
ccc
(2000a)*
Rabbit
NZW
F
5
0.0
13.9
32.3
76.3
Drinking
water
Gestation
days 6-29
Reproductive and
developmental
parameters
76.3
(developmental)
-
Not modeled
—
CCC
(2000b)*
Rabbit
NZW
F
25
0
1.4
13.4
Drinking
Water
Gestation
days 6-29
Clinical sign, gross
lesions, reproductive
and developmental
endpoints
55
(developmental)
-
Not modeled
-
35.6
55.3
13.4
(maternal)
35.3
(reduced
corrected
maternal body
weight gain
gestation days
6-29)
50
BMDS software
failed
* The NOAEL and LOAEL values listed are for reproductive or developmental effects.
** The NOAEL and LOAEL values were the same for developmental effects in either vehicle. The LOAEL for maternal toxicity was 25 mg/kg-day for the
aqueous vehicle (10% Emulphor). The NOAEL and LOAEL for maternal toxicity using the corn oil vehicle were 25 mg/kg-day and 50 mg/kg-day,
respectively. BMD modeling was performed on aqueous vehicle data only.
^ BMD and BMDL10 calculated using duration adjusted doses
Abbreviations: FEL, Frank effect level; SD, Sprague-Dawley; NZW, New Zealand White
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VIII - 13
February 20, 2003
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c. Longer-term Health Advisory
Two rodent oral exposure studies conducted by NTP (1987) were considered for
derivation of the Longer-term HA for bromodichloromethane. In addition, eight reproductive
studies were considered. These studies are summarized in Table VIII-3 below.
NTP (1987) administered bromodichloromethane by gavage to F344/N rats for 5
days/week for 13 weeks at dose levels ranging from 19 to 300 mg/kg-day. Based on decreased
weight gain, this study identified a NOAEL of 75 mg/kg-day and a LOAEL of 150 mg/kg-day.
Treatment-related hepatic and renal lesions were observed only at the high dose. In a similar
study, NTP (1987) administered bromodichloromethane by gavage to B6C3F, mice for 5
days/week for 13 weeks at dose levels ranging from 6.25 to 100 mg/kg-day for males and from 25
to 400 mg/kg-day for females. This study identified a NOAEL of 50 mg/kg-day and a LOAEL of
100 mg/kg-day based on histologic alterations in the kidney (focal necrosis of the proximal renal
tubular epithelium and nephrosis) of male mice. BMD analysis using the BMDS program
identified duration-adjusted BMD values of 63 and 75 mg/kg-day for focal necrosis of renal
tubular epithelium in males and vacuolated cytoplasm in the liver of females, respectively. The
corresponding BMDL10 values for these renal and hepatic effects were 35 and 47 mg/kg-day,
respectively.
Eight reproductive or developmental studies (Ruddick et al., 1983; Narotsky et al., 1997;
Bielmeier et al. 2001; CCC, 2000a,b,c,d; CCC, 2002)were also considered for derivation of the
Longer-term HA. CCC (2002) identified a LOAEL of 150 ppm (approximately 11.6 to 40.2
mg/kg-day) for delayed sexual maturation in F, male rats in a two-generation study of
bromodichloromethane administered in drinking water. The LOAEL for parental effects in the
study was also 150 ppm, based on decreased body weight and body weight gain in F, females and
F, males and females. Bielmeier et al. (2001) examined the occurrence of full litter resorption in
F344 rats treated with 0, 75 or 100 mg/kg-day bromodichloromethane by aqueous gavage on
gestation day 9. The LOAEL for this effect was 75 mg/kg-day. Narotsky et al. (1997) evaluated
the same endpoint in F344 rats administered 0, 25, 50, or 75 mg/kg-day on gestation days 6
through 15. Full litter resorption was observed at 50 and 75 mg/kg-day. The NOAEL and
LOAEL in this study were thus identified as 25 and 50 mg/kg-day, respectively. When data from
these studies were analyzed using the BMD approach, BMD values of 48 and 23 mg/kg-day were
obtained for the Narotsky et al. (1997) and Bielmeier et al. (2001) studies, respectively. The
higher value from the Narotsky et al. (1997) study was considered the more reliable estimate of
the BMD because it was based on response data that included lower doses, one of which was an
apparent NOAEL. The BMDL10 calculated from the Narotsky et al. (1997) data was 30 mg/kg-
day. Studies conducted by NTP (1998) did not detect reproductive or developmental toxicity at
doses up to 116 mg/kg-day. Two studies conducted in New Zealand White rabbits did not detect
developmental effects at doses up to 55 and 76 mg/kg-day, respectively (CCC, 2000a,b).
Three additional studies in rats identified developmental effects that occurred at dose
levels that also resulted in maternal toxicity. In a range-finding study (CCC, 2000c), F,
generation pups exposed to bromodichloromethane via lactation and possibly by consumption of
Draft - Do Not Cite or Quote
VIII - 14
February 20, 2003
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water supplied to the dams exhibited reduced body weights and body weight gains. Biologically
meaningful daily doses could not be established in this experiment; therefore, the concentration-
based NOAEL and LOAEL values are 50 ppm and 150 ppm. based on reduced body weight and
body weight gain in the F, pups. In a subsequent developmental study, CCC (2000d) identified
NOAEL and LOAEL values of 45 and 82 mg/kg-day, respectively, based on decreased number of
ossification sites per fetus for the forelimb phalanges and hindlimb metatarsals and phalanges.
These reversible developmental delays occurred at doses which also resulted in maternal.
Endpoints from these studies were not modeled because other effects were observed at lower
doses. Ruddick et al. (1983) observed an increased incidence of sternebral aberrations in the pups
of Sprague-Dawley rats administered bromodichloromethane in corn oil by gavage. Statistical
analysis of the published data indicated that the NOAEL and LOAEL for this effect were 100
mg/kg-day and 200 mg/kg-day, respectively. The lowest dose tested was 50 mg/kg-day. The
BMD and BMDL10 obtained for this study were 27 and 15 mg/kg-day, respectively. However,
examination of the modeling output indicated that none of the available models fit the data well in
the low-dose region of the curve. Therefore, the reliability of these values is questionable.
The Narotsky et al. (1997) and CCC (2000d) studies identified BMDL10 values of 10 and
18 mg/kg-day based on reduced maternal body weight gain. The CCC data were considered the
most relevant since they were obtained from a drinking water study which utilized concentrations
that resulted in daily doses well below those used in the Narotsky study. The NTP (1987) and the
Narotsky et al. (1997) studies provided similar, but higher, BMDL10 values, based on
reproductive and histopathological endpoints. The NTP (1987) study utilized
bromodichloromethane doses as low as 6.3 mg/kg-day, in contrast to the reproductive study
conducted by Narotsky et al. (1997) in which the lowest dose was 25 mg/kg-day. The NTP
(1987) data thus provide more information about the shape of the dose-response curve in the
region of interest. The BMD data for focal necrosis of renal tubular epithelium and reduced body
weight gain in pregnant female rats were thus selected as the most reliable basis for determining
the Longer-term HA. Using the lower of the two values, the duration-adjusted BMDL10 of 18
mg/kg-day for reduced maternal body weight gain, the Longer-term HA for a 10 kg child is
calculated according to the following equation:
(18 mg/kg-day) (10 kg)
Longer-term HA = = 0.60 mg/L (rounded to 0.6 mg/L)
(300) (1 L/day)
where:
18 mg/kg-day = BMDL10 based on decreased maternal body weight gain on
gestation days 6-9 in pregnant female rats administered
bromodichloromethane in the drinking water.
10 kg = Assumed body weight of a child
Draft - Do Not Cite or Quote
VIII - 15
February 20, 2003
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300 = Uncertainty factor based on NAS/OW guidelines. This
value includes a factor of 10 to protect sensitive human
populations and a factor of 10 for extrapolation from
animals to humans, and a factor of 3 to account for
uncertainty regarding possible reproductive effects of
bromodichloromethane in humans
1 L/day = Assumed water consumption of a 10-kg child
The Longer-term HA for adults would be calculated as follows:
(18 mg/kg-day) (70 kg)
Longer-term HA = = 2.1 mg/L (rounded to 2 mg/L)
(300) (2 L/day)
where:
18 mg/kg-day
70 kg
300
= BMDL10 based on reduced body weight gain in female rats
administered bromodichloromethane in the drinking water
= Assumed body weight of an adult
= Uncertainty factor based on NAS/OW guidelines. This
value includes a factor of 10 to protect sensitive human
populations and a factor of 10 for extrapolation from
animals to humans, and a factor of 3 to account for
database limitations and uncertainty related to potential
reproductive effects of bromodichloromethane in humans
2 L/day = Assumed water consumption of a 70-kg adult
For purposes of comparison , a Longer-term HA derived using the conventional
NOAEL/LOAEL approach would be based on the study conducted by CCC (2000d). This study
identified a NOAEL of 18.4 mg/kg-day and a LOAEL of 45 mg/kg-day based on reduced body
weight gain on gestation days 6 to 7. Using the NOAEL of 18.4 mg/kg-day, and assuming
drinking water ingestion of 1 L/day and an uncertainty factor of 300 (including factors of 10 for
interspecies extrapolation and protection of susceptible populations, and a factor of 3 for database
limitations and uncertainty regarding potential reproductive effects in humans), the Longer-term
HA for a 10 kg child would be 0.6 mg/L. The Longer-term HA for a 70 kg adult consuming 2
L/day would be 2 mg/kg-day.
Draft - Do Not Cite or Quote
VIII - 16
February 20, 2003
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Table VTTT-3 Summary of Candidate Studies for Derivation of the Longer-term HA for Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL1C
(mg/kg-day) {
NTP
(1987)
Rat
F344/N
M, F
10
0
19
38
75
150
300
Gavage
(oil)
13 weeks
(5 d/wk)
Body weight,
clinical signs,
histology
75
150
(decreased
weight gain)
(hepatic and
renal lesions at
300)
Not modeled
NTP (1987)
Mouse
B6C3Fj
M, F
10
Male
0
6.3
13
25
50
100
Female
0
25
50
100
200
400
Gavage
(oil)
13 weeks
(5 d/wk)
Body weight,
clinical signs,
histology
50
100
(renal lesions)
63
35
(focal necrosis of
renal tubular
epithelium in
males)
75
47
(Hepatic
vacuolated
cytoplasm in
females)
Ruddick et
al. (1983)*
Rat
SD
F
9-
14
0
50
100
200
Gavage
(oil)
Gestation days
6-15
Body and organ
weights;
maternal serum
chemistry;
hematology, and
histopathology;
developmental
parameters
100
200
(increased
incidence of
sternebral
variations)
27
15
(increased
incidence of
sternebral
variations)
Narotsky et
al.
(1997)*
Rat
F344
F
12-
14
0
25
50
75
Gavage
(oil)
(aq)**
Gestation days
6-15
Body weight,
clinical signs,
developmental
parameters
25
(developmental)
50
(full-litter
resorption)
48
30
(full-litter
resorption)
(maternal)
25
(reduced
maternal body
weight gain
gestation days 6-
8, aq. vehicle)
18
10
(reduced
maternal body
weight gain,
gestation days 6-
8, aq. vehicle)
Draft - Do Not Cite or Quote
VIII - 17 February 20, 2003
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Table VIII-3 (cont.)
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day) {
Bielmeier et
al. (2001)*
Rat
F344
F
8-
11
0
75
100
Gavage
(aq)
Gestation day
9
Full litter
resorption;
hormone
profiles
75
(full-litter
resorption)
23
4.2
(full-litter
resorption)
CCC
(2000c)*
Rat
SD
M, F
10
0 ppm
50 ppm
150 ppm
450 ppm
1350
ppm
Drinking
water
Males
64 days
Females
74 days
Reproductive/
developmental
parameters
Developmental
50 ppm
Parental
50 ppm
Developmental
150 ppm-
Parental
50 ppm
Not modeled
CCC
(2000d)*
Rat
SD
F
25
0.0
2.2
18.4
45.0
82.0
Drinking
water
Gestation days
6-21
Reproductive/
developmental
parameters
45.3
(developmental)
82.0
(reduced number
of ossification
sites in
phalanges or
metatarsals
occurring with
maternal
toxicity)
Not modeled
18.4
(maternal)
45
(reduced
maternal body
weight gain
gestation days 6-
7)
23
18
(reduced
maternal body
weight gain
gestation days 6-
9; see comments
in text)
CCC (2002)*
Rat
SD
M, F
30
0 ppm
50 ppm
150 ppm
450 ppm
Drinking
water
Two
generations
Reproductive
parameters
50 ppm
(offspring)
150 ppm
(delayed sexual
maturation in ¥1
males)
Not modeled
50 ppm
(parental)
150 ppm
(Reduced body
wt and body wt
gain in F0
females and ¥1
males and
females)
Draft - Do Not Cite or Quote
VIII - 18 February 20, 2003
-------
Table VIII-3 (cont.)
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day) {
ccc
(2000a)*
Rabbit
NZW
F
5
0
4.9
13.9
32.3
76.3
Drinking
Water
Gestation day
6 to 29
Body weight,
clinical signs,
reproductive and
developmental
parameters
76.3
(developmental)
-
Not modeled
CCC
(2000b)*
Rabbit
NZW
F
25
0
1.4
13.4
Drinking
Water
Gestation days
6-29
Clinical sign,
gross lesions,
reproductive and
55.3
(developmental)
"
Not modeled
~
35.6
55.3
developmental
endpoints
13.4
(maternal)
35.3
(reduced
maternal body
weight gain
50
(developmental)
BMD software
failed
} BMDL10 value was derived using duration-adjusted doses.
¦f Modeled using Crump Benchmark Dose Software
* The NOAEL and LOAEL values listed are for reproductive/developmental effects.
** The developmental NOAEL and LOAEL values were the same in either vehicle. The LOAELs for maternal toxicity were 25 mg/kg-day and 50 mg/kg-
day for the aqueous and corn oil vehicles respectively. BMD modeling was performed on aqueous vehicle data only.
Abbreviations: NA, Not available; SD, Sprague-Dawley; NZW, New Zealand White
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VIII - 19 February 20, 2003
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d. Reference dose, Drinking Water Equivalent Level and Lifetime Health Advisory
This section reports the existing RfD value for bromodichloromethane and describes the
derivation of the RfD for this compound. This section also describes the calculation of Drinking
Water Equivalent Level and Lifetime Health Advisory values which require the RfD as input. For
this document, new and existing studies were reviewed and appropriate candidate data were
selected for benchmark dose (BMD) modeling. The results of BMD modeling were used in
conjunction with appropriate uncertainty factors to calculate the RfD. A comparison of the RfD
derived using the BMD approach to the results obtained using the conventional NOAEL/LOAEL
approach is also provided.
Description of the Existing RfD
The existing RfD for bromodichloromethane is 0.02 mg/kg-day (IRIS, 1993a). This value
was derived using a duration-adjusted LOAEL of 17.9 mg/kg-day identified for renal cytomegaly
in B6C3F, mice administered bromodichloromethane by corn oil gavage for 5 days/week for 102
weeks (NTP, 1987). An uncertainty factor of 100 was used to account for extrapolation from
animal data and for protection of sensitive human subpopulations. An additional factor of 10 was
used because the RfD was based on a LOAEL (although it was considered minimally adverse) and
to account for lack of reproductive data.
Identification of Candidate Studies for Derivation of the RfD
Several studies of chronic duration were considered for derivation of the RfD for
bromodichloromethane. These studies are summarized in Table VIII-4 below. NTP (1987)
administered bromodichloromethane to F344/N rats by gavage in corn oil at doses of 50 or 100
mg/kg-day for 5 day/week for 102 weeks. This study identified a LOAEL of 50 mg/kg-day based
on histologic alterations in the liver and kidney. In a similar study, NTP (1987) administered
bromodichloromethane by gavage in corn oil to B6C3F, mice for 5 days/week for 102 weeks at
dose levels of 25 or 50 mg/kg-day for males and 75 or 150 mg/kg-day for females. Based on
histologic alterations in the liver, kidney, and thyroid of male mice, this study identified a LOAEL
of 25 mg/kg-day, which is consistent with the value identified in the rat study. In a third study,
Tobe et al. (1982) administered microencapsulated bromodichloromethane to Wistar rats in the
diet at dose levels ranging from 6 to 168 mg/kg-day. Histologic data for the animals exposed to
bromodichloromethane were reported by Aida et al. (1992b). This study identified a LOAEL for
male rats of 6 mg/kg-day on the basis of histopathologic changes in the liver.
Ten reproductive and/or developmental toxicity studies (Ruddick et al., 1983; Klinefelter
et al., 1995; Narotsky et al., 1997; NTP, 1998; Bielmeier et al. 2001; CCC, 2000a,b,c,d; CCC,
2002) were considered for derivation of the RfD in addition to the chronic studies. The
investigations of Ruddick et al. (1983), Klinefelter et al. (1995), and Narotsky et al. (1997)
identified NOAELs or LOAELs in rats that were substantially higher than the LOAEL identified
by Aida et al. (1992b) (Table VIII-4 below). The studies conducted by NTP (1998) did not
observe developmental or reproductive effects at doses up to 116 mg/kg-day. The study by
Bielmeier et al. (2001) identified a free-standing LOAEL of 75 mg/kg-day in F344 rats. The
studies conducted by CCC (2000a,b) identified NOAELs of 55 and 76 mg/kg-day, respectively,
Draft - Do Not Cite or Quote
VIII - 20
February 20, 2003
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for developmental effects in New Zealand White rabbits. The study conducted in rats by CCC
(2000c) identified concentration-based NOAEL and LOAEL values of 50 ppm and 150 ppm for
reduced body weight and body weight gain in F, pups. The study conducted in rats by CCC
(2000d) identified NOAEL and LOAEL values of 45 mg/kg-day and 82 mg/kg-day, respectively
on the basis of decreased ossification sites per fetus per litter in the forelimb and hindlimb. Low-
range LOAEL values for maternal toxicity ranged from 13 to 25 mg/kg-day (CCC, 2000b; CCC
2000d; Narotsky et al., 1997). CCC (2002) identified a LOAEL of 150 ppm (approximately 11.6
to 40.2 mg/kg-day) for delayed sexual maturation in F, male rats in a two-generation study of
bromodichloromethane administered in drinking water. The LOAEL for parental effects in the
study was also 150 ppm, based on decreased body weight and body weight gain in F, females and
F, males and females. Since these studies identified NOAEL and/or LOAEL values substantially
higher than that identified by Aida et al. (1992b), they were not further considered for derivation
of the RfD.
Method of Analysis
Selected data from the candidate studies were analyzed using the benchmark dose (BMD)
modeling approach. Initially, data sets for potentially sensitive endpoints were selected as
described in U.S. EPA (1998b) and analyzed using the Crump Benchmark Dose Modeling
Software (K. S. Crump, Inc.). Results of this analysis are summarized in Table VIII-5. Following
the release of Version 1.2 of the BMDS program (U.S. EPA, 2000a), a subset of the most
sensitive endpoints identified using the Crump software was reanalyzed in accordance with
proposed U.S. EPA (2000b) recommendations. An advantage of analysis with the BMDS
software is that several additional models are available to fit the data. The results of the analysis
with the BMDS software are included in Table VIII-4.
Choice of Principal Study and Critical Effect for the RfD
Three data sets for histopathological effects in liver were analyzed using the BMDS
software (Table VIII-4). BMD modeling identified several endpoints with BMD values lower
than the conventionally determined LOAEL of 6 mg/kg-day (Aida, 1992b). The lowest BMD
values were obtained for fatty degeneration (1.9 mg/kg-day) in male rats (Aida et al., 1992b) and
for kidney cytomegaly (2.0 mg/kg-day) in male mice (NTP, 1987). Comparably low BMD values
were also obtained for granulomas observed in the liver of male rats (2.1 mg/kg-day) and for fatty
degeneration in the liver of female rats (3.1 mg/kg-day) in the study conducted by Aida et al.
(1992b). In contrast, the BMD values calculated for endpoints examined in other studies were
approximately 10- to 20-fold higher.
Draft - Do Not Cite or Quote
VIII-21
February 20, 2003
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Table VIII-4 Summary of Candidate Studies for Derivation of the RfD for Bromodichloromethane
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL1C
(mg/kg-day) ¦
NTP
(1987)
Rat
F344/N
M, F
50
0
50
100
Gavage
(oil)
102 weeks
(5 d/wk)
Body weight,
clinical signs,
gross necropsy,
histology
--
50
(lesions of
kidney and liver)
-
36.5°
(liver necrosis in
male rats)
NTP
(1987)
Mouse
B6C3Fj
M, F
50
0
25
50
Gavage
(oil)
102 weeks
(5 d/wk)
Body weight,
clinical signs,
gross necropsy,
histology
--
25
(lesions of liver,
kidney, and
thyroid)
2.0
1.5
(kidney
cytomegaly in
male mice)
Aida et al.
(1992b)
Rat
Wistar
M, F
40
Male
0
6
26
Diet
24 months
Body weight,
clinical signs,
serum
biochemistry,
--
6
(liver fatty
degeneration and
granuloma)
3.1
2.1
(fatty
degeneration in
liver of females)
138
Female
0
8
32
168
gross necropsy,
histology
1.9
0.8
(fatty
degeneration in
liver of males)
2.1
1.4
(Granulomas in
livers of males)
Ruddick et
al. (1983)b
Rat
SD
F
9-
14
0 ppm
50 ppm
150 ppm
450 ppm
1350
ppm
Gavage
(oil)
Gestation days
6-15
Body and organ
weights;
maternal serum
chemistry;
hematology, and
histopathology;
developmental
parameters
100
200
(sternebral
variations)
27
15
(sternebral
variations)
Narotsky et
al. (1997)b
Rat
F344
13-
14
0
75
100
Gavage
(oil)
(water)
Gestation days
6-15
Body weight,
clinical signs,
developmental
parameters
25
(developmental)
50
(full-litter
resorption)
48
30
(full litter
resorption)
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Table VIII-4 (cont.)
Klinefelter
etal. (1995)b
Rat
F344
M
7
0
22
39
Drinking
water
52 weeks
Body and organ
weights, gross
necropsy,
histology, sperm
motion
parameters
22
39
(decreased sperm
velocities)
__d
Bielmeier et
al. (200 l)b
Rat
F344
F
8-
11
0
75
100
Gavage
(aq)
Gestation day
9
Full litter
resorption,
hormone
profiles, body
weight
75
(full-litter
resorption)
23
4.2
(full-litter
resorption)
CCC
(2000a)b
Rabbit
NZW
F
5
0
4.9
13.9
32.3
76.3
Drinking
Water
Gestation days
6-29
Body wt.,
clinical signs,
reproductive
developmental
parameters
76
e
CCC
(2000b)b
Rabbit
NZW
F
25
0
1.4
13
36
55
Drinking
water
Gestation days
6-29
Maternal feed
and water
intake, body
wt.; gross
lesions; uterine
weight, no.
implantation
55
(developmental)
e
sites, uterine
contents, and no.
corpora; Fetal
wt., gross ext.
alterations, skel.
alterations, sex,
visceral
alterations
13.4
(maternal)
35.3
(reduced
maternal body
weight gain)
50
(maternal)
BMDS software
failed
CCC
(2000c)b
Rat
SD
M,F
10
0 ppm
50 ppm
150 ppm
450 ppm
1350
ppm
Drinking
water
Males
64 days
Females
74 days
Reproductive/
developmental
parameters
50 ppm
150 ppm
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Table VIII-4 (cont.)
CCC
(2000d)b
Rat
SD
F
25
0.0
2.2
18.4
45.0
82.0
Drinking
water
Gestation days
6-21
Reproductive/
developmental
parameters
45
(developmental)
82
(reduced no. of
ossification sites
in phalanges or
metatarsals
occurring with
maternal
toxicity)
(developmental)
18.4
(maternal)
45
(reduced
maternal body
weight gain)
23
(maternal)
18
(reduced
maternal body
weight gain)
CCC
(2002)
Rat
SD
M, F
30
0 ppm
50 ppm
150 ppm
450 ppm
Drinking
water
Two
generations
Reproductive/
developmental
parameters
50
(offspring)
150
(delayed sexual
maturation in ¥1
males)
Not modeled
--
50
(parental)
150
(decrecreased
body weight and
body weight gain
in F0 females and
Fj males and
females)
Not modeled
a BMDL10 values were derived using duration-adjusted doses.
b Ruddick et al. (1983); Klinefelter et al. (1995), Narotsky et al. (1997), Bielmeier et al. (2001), and CCC (2000a-d)are included in this table because they are
reproductive and/or developmental studies. The NOAEL, LOAEL, BMD, and BMDL10 values listed are for reproductive and/or developmental endpoints.
c Data modeled using Crump BMD software
d No histopathological abnormalities were noted in this study, and similar effects on sperm velocity were not observed in NTP (1998); therefore, data for
sperm velocity were not modeled
e Data were not modeled since effects occurred at higher doses than other candidate endpoints
- Indicates that data were not modeled
Abbreviations: NZW, New Zealand White; SD, Sprague-Dawley
NOTE: The short-term reproductive and developmental toxicity study conducted by NTP (1998) was not included in this table because no developmental or
reproductive effects were noted at dose levels ranging from 67 to 126 mg/kg-day.
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Table VI11-5 Summary of Preliminary BMD Modeling Results for the
Bromodichloromethane RfD
Study
Endpoint Modeled
BMDL10 (mg/kg-day) *
Subchronic NTP (1987)
mouse study
Focal necrosis of renal tubular epithelium in males
34
Vacuolated hepatocytes in females
64
Chronic NTP (1987) rat
study
Kidney cytomegaly in males
No acceptable fit
Liver necrosis in males
36.5
Liver fatty metamorphosis in males
No acceptable fit
Clear cell changes in liver of females
No acceptable fit
Chronic NTP (1987) mouse
study
Kidney cytomegaly in males
0.96
Liver fatty metamorphosis in males
7.5
Thyroid follicular cell hyperplasia in females
15
Chronic Aida et al. (1992b)
rat study
Fatty degeneration in liver of males
2.38
Granulomas in liver of males
4.5
Fatty degeneration in liver of females
1.20
Granulomas in liver of females
No acceptable fit
* BMD modeling conducted on duration-adjusted doses using the Cramp BMD Software (K. S. Cramp, Inc.).
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The chronic study conducted by Aida et al. (1992b) was selected for derivation of the
RfD. The lowest dose utilized in this study was 6 mg/kg-day (in contrast to low doses of 22 to 75
mg/kg-day utilized in other candidate studies), which provides some information on the shape of
the dose-response curve in the region of interest. The lowest BMD (1.9 mg/kg-day) was obtained
for fatty degeneration in the liver of male mice. The corresponding BMDL10 for this endpoint
was 0.8 mg/kg-day. The incidence of this lesion was strongly dose-dependent, with incidences of
0/24, 5/14, 12/13, and 19/19 observed at the doses of 0, 6, 25, and 138 mg/kg-day, respectively.
The occurrence of this lesion in rats treated with bromodichloromethane is consistent with current
understanding of the mode of action of brominated trihalomethanes. This endpoint was therefore
selected to derive the RfD for bromodichloromethane.
Derivation of the RfD
The BMDL10 calculated for fatty degeneration in the liver of male rats in the chronic rat
study conducted by Aida et al. (1992b) was selected as the most appropriate basis for derivation
of the RfD for bromodichloromethane. The RfD is calculated according to the following
equation:
(0.8 mg/kg-day)
RfD = = 0.003 mg/kg-day (3 |ig/L)
(300)
where:
0.8 mg/kg-day = Duration-adjusted BMDL10 based on fatty degeneration of the
liver in male rats
300 = Uncertainty factor based on NAS/OW guidelines. This value
includes a factor of 10 to account for intrahuman variability,
and a factor of 10 for interspecies variability, and a factor of 3
to account for database deficiencies, including lack of a multi-
generation reproductive toxicity study and database limitations
and uncertainty related to possible human reproductive effects
suggested (causality can not be established from available
data) by epidemiological studies.
A composite UF of 300 was used. The standard factors of 10 were used for interspecies
extrapolation and for protection of sensitive subpopulations. An additional factor of 3 was used
to account for database deficiencies related to possible reproductive or developmental effects in
humans. Use of an additional uncertainty factor of 3 is supported by findings in epidemiological
studies (Waller et al., 1998; King et al., 2000) which suggest potential associations between
bromodichloromethane exposure via drinking water and adverse pregnancy outcomes. Although
the results of the epidemiological studies can not establish that bromodichloromethane caused the
observed effects, they do raise significant concern for potential reproductive effects in exposed
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human populations, and the inclusion of an additional uncertainty factor is thus considered
appropriate for protection of human health.
The DWEL for bromodichloromethane is calculated as follows:
(0.003 mg/kg-day) (70 kg)
DWEL = = 0.100 mg/L (100 |ig/L)
2 L/day
where:
0.003 mg/kg-day = RfD for bromodichloromethane
70 kg = Assumed weight of an adult
2 L/day = Assumed water consumption by a 70-kg adult
Lifetime Health Advisory
The Lifetime Health Advisory (HA) represents that portion of an individual's total
exposure that is attributed to drinking water and is considered protective of noncarcinogenie
health effects over a lifetime of exposure. Bromodichloromethane has been categorized with
respect to carcinogenic potential as Group B2: Probable human carcinogen (IRIS, 1993a).
Therefore, in accordance with U.S. EPA Policy, a Lifetime HA is not recommended.
Alternative Approach for Derivation of the RfD
Use of the conventional NOAEL/LOAEL approach represents an alternative means for
deriving the RfD and DWEL. Aida et al. (1992b) identified a LOAEL of 6 mg/kg-day in male
rats on the basis of histopathological changes in the liver. Using this value and a composite
uncertainty factor of 3,000 (including factors of 10 for interspecies extrapolation, protection of
sensitive subpopulations, and use of a LOAEL, and a factor of 3 for database limitations and
uncertainty regarding potential reproductive effects in humans), the RfD derived using the
conventional approach is 0.002 mg/kg-day. Assuming a body weight of 70 kg and drinking water
ingestion of 2 L/day, the corresponding DWEL is 0.07 mg/L (70 |ig/L).
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2. Carcinogenic Effects
a. Categorization of Carcinogenic Potential
Previous Evaluations
The Carcinogenic Risk Assessment Verification Endeavor (CRAVE) group of the
U.S. EPA reviewed the available evidence on the carcinogenicity of bromodichloromethane and
assigned it to Group B2: probable human carcinogen (IRIS, 1993a). Assignment to this category
is appropriate for chemicals where there are no or inadequate human data, but which have
sufficient animal data to indicate carcinogenic potential.
IARC has recently re-evaluated the carcinogenic potential of bromodichloromethane
(IARC 1999a). IARC concluded that there is sufficient evidence of carcinogenicity for
bromodichloromethane in experimental animals, but inadequate evidence in humans. Thus, IARC
classified bromodichloromethane as a Group 2B carcinogen: possibly carcinogenic to humans.
Categorization of Carcinogenic Potential Under the Proposed 1999 Cancer Guidelines
Cancer Hazard Summary
Under the proposed guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999)
bromodichloromethane is likely to be carcinogenic to humans by all routes. This descriptor is
appropriate when the available tumor data and other key data are adequate to demonstrate
carcinogenic potential to humans. This finding is based on the weight of experimental evidence in
animal models which shows carcinogenicity by modes of action that are relevant to humans.
Supporting Information for Cancer Hazard Assessment
Human Data
The information on the carcinogenicity of bromodichloromethane from human studies is
inadequate. There are no epidemiological data specifically relating increased incidence of cancer
to exposure to bromodichloromethane. There are equivocal epidemiological data describing a
weak association of chlorinated drinking water exposures with increased incidences of bladder,
rectal, and colon cancer. U.S. EPA has determined that these studies cannot attribute the
observed effects to a single compound, as chlorinated water contains numerous other disinfection
byproducts that are potentially carcinogenic.
Animal Data
The carcinogenicity of bromodichloromethane in male and female animals has been
investigated in a well-designed and conducted corn oil gavage study conducted in rats and mice, a
dietary exposure study in rats, and two drinking water studies in rats. Additional data are
available from a study in which male Strain A mice were exposed to bromodichloromethane by
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intraperitoneal injection. No data are available on the carcinogenic potential of
bromodichloromethane administered via the inhalation or dermal routes.
In the corn oil gavage study (NTP, 1987), statistically significant increases were observed
in the incidences of neoplasms of the large intestine and kidney in male and female rats, the kidney
in male mice, and liver in female mice. The neoplasms observed in the large intestine and kidney
are considered rare neoplasms based on historical control data for the tested strains. In the
feeding study (Aida et al., 1992b), exposure to microencapsulated bromodichloromethane did not
result in statistically significant increases in any tumor type. Observed neoplastic lesions included
three cholangiocarcinomas and two hepatocellular adenomas in high dose females, one
hepatocellular adenoma in a control female, one cholangiosarcoma in a high dose male, and one
hepatocellular carcinoma each in a low- and a high dose male. In the drinking water study
conducted by Tumasonis et al. (1985), hepatic neoplastic nodules, hepatic adenofibrosis, and
lymphosarcoma were significantly increased in female rats. No significant increase in the
occurrence of any tumor type was observed in male rats. Renal adenoma or adenocarcinoma
were noted in two males and one female treated with bromodichloromethane, while neither tumor
type was reported in the control group. In the drinking water study conducted by George et al.
(2002), the prevalence of neoplastic lesions in the liver was significantly increased only at the
lowest administered dose. Intraperitoneal injection of Strain A mice with bromodichloromethane
resulted in an apparent increase in the number of pulmonary adenomas per animals, although the
response did not reach statistical significance in any dose group.
Structural Analogue Data
Trihalomethanes structurally related to bromodichloromethane have shown varying
degrees of carcinogenic potential in rodents. Chloroform, the most extensively characterized
trihalomethane, is reported to be carcinogenic at high doses in several chronic animal bioassays,
with significant increases in the incidence of liver tumors in male and female mice and significant
increases in the incidence of kidney tumors in male rats and mice (U.S. EPA, 2001). The
occurrence of tumors in animals exposed to chloroform is demonstrably species-, strain-, and
gender-specific, and has only been observed under dose conditions that caused cytotoxicity and
regenerative cell proliferation in the target organ. The cancer database for structurally-related
brominated trihalomethanes is more limited, but includes well-conducted studies performed by the
National Toxicology Program. In a two-year corn oil gavage study of bromoform, NTP (1989a)
found clear evidence for carcinogenicity in female rats and some evidence of carcinogenicity in
male rats based on occurrence of tumors of the large intestine (adenomatous polyps or
adenocarcinoma). In a two-year corn oil gavage study of dibromochloromethane, NTP (1985)
determined that there was some evidence of carcinogenicity in female mice and equivocal
evidence of carcinogenicity in male mice, based on the occurrence of hepatocellular adenomas and
carcinomas. Other, less well-documented, oral exposure studies (Tobe et al., 1982; Kurokawa,
1987; Voronin et al., 1987) found no evidence for carcinogenicity of bromoform or
dibromochloromethane.
Other Key Data
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Bromodichloromethane is formed as a byproduct of drinking water disinfection with
chlorine. Exposure to bromodichloromethane may occur via ingestion of tap water, via dermal
contact during showering or bathing, or by inhalation of bromodichloromethane volatilized during
household activities. Absorption of single oral doses appears to be extensive.
Bromodichloromethane is rapidly metabolized and eliminated predominately as expired volatiles,
C02, or CO. Only a small amount (less than 10%) is eliminated in urine or in feces. No
comprehensive tissue data are available regarding the bioaccumulation or retention of
bromodichloromethane following repeated exposure. However, because of the rapid metabolism
and excretion of bromodichloromethane, marked accumulation and retention is not expected.
Bromodichloromethane itself is not directly reactive with DNA. Metabolism to reactive
species is a prerequisite for toxicity, as inferred from metabolic induction and inhibition studies.
In vitro and in vivo studies of the mutagenic and genotoxic potential of bromodichloromethane
have yielded both positive and negative results. Synthesis of the overall weight of evidence from
these studies is complicated by the use of a variety of testing protocols, different strains of test
organisms, different activating systems, different dose levels, different exposure methods (gas
versus liquid) and, in some cases, problems due to evaporation of the test chemical. However,
because a majority of studies yielded positive results, bromodichloromethane is considered to be
at least weakly mutagenic and genotoxic. Recent studies conducted with strains of Salmonella
engineered to express rat theta-class glutathione »Y-transferase suggest that mutagenicity of the
brominated trihalomethanes may be mediated by glutathione conjugation.
Mode of Action
The mode of action for tumor induction by bromodichloromethane has not been clearly
elucidated and may involve contributions from multiple bioactivation pathways. In each case,
toxicity is believed to result from interaction of reactive metabolites with cellular macromolecules.
Proposed bioactivation pathways for bromodichloromethane include: 1) production of reactive
dihalocarbonyls by oxidative metabolism; 2) production of reactive dihalomethyl radicals by
oxidative metabolism; and 3) formation of DNA-reactive species via a glutathione-dependent
pathway. The relative contribution of each pathway to tumor induction by
bromodichloromethane has not been characterized. It is possible that only the latter two
processes lead to DNA damage in vivo, because the highly reactive dihalocarbonyl intermediate
may not survive long enough to enter the nucleus and react with DNA. For this reason,
cytotoxicity may be the primary consequence of the oxidative pathway. Cytotoxicity coupled
with regenerative hyperplasia is considered the primary mode of action for tumor formation
following exposure to high concentrations of chloroform, a structurally-related trihalomethane
which has low genotoxic potential. Data to support a similar primary mode of action for tumor
development in liver, kidney, and large intestine are currently lacking for bromodichloromethane.
In the absence of such information, combined with a positive weight-of-evidence evaluation for
genotoxicity, the mode of action for tumor development is assumed to be a linear process. The
processes leading to tumor formation in animals are expected to be relevant to humans.
Conclusion
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Under the proposed guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999)
bromodichloromethane is likely to be carcinogenic to humans by the oral route. This weight-of-
evidence evaluation is based on 1) observations of tumors in animals treated by oral pathways; 2)
lack of epidemiological data specific to bromodichloromethane and equivocal data for drinking
water drinking water exposures that cannot reliably be attributed to bromodichloromethane
among multiple other disinfection byproducts; 3) positive results for a majority of the available
genotoxicity and mutagenicity tests; and 4) metabolism and mode of action that are reasonably
expected to be comparable across species. Although no cancer data exist for exposures via the
dermal or inhalation pathways, the weight-of-evidence conclusion is considered to be applicable
to these pathways as well. The finding for inhalation is based on the observation that patterns of
metabolizing enzyme activity in male rats are similar for exposure via the inhalation and gavage
routes. Bromodichloromethane absorbed through the skin is expected to be metabolized and
cause toxicity in much the same way as bromodichloromethane absorbed by the oral and
inhalation routes.
b. Choice of Study for Quantification of Carcinogenic Risk
In accordance with the Proposed 1999 Cancer Guidelines (U.S. EPA, 1999),
quantification of cancer risk is appropriate for compounds categorized as likely to be
carcinogenic to humans. Five oral exposure studies were available for quantification of the
carcinogenic risk associated with exposure to bromodichloromethane. Detailed summaries of
these studies are provided in Section V.G. 1. The two-year study conducted by NTP (1987) in
rats and mice was selected for quantification of carcinogenic effects associated with exposure to
bromodichloromethane. Selection of this study was based on significantly increased incidence of
several tumor types, monotonic dose response curves, and comprehensive documentation of study
design and results.
In the NTP (1987) study, groups of male and female F344/N rats (50/sex/dose) received
0, 50, or 100 mg/kg-day gavage doses of bromodichloromethane in corn oil. The doses were
administered 5 days/week for 102 weeks. In a similar experiment, groups of male and female
B6C3FJ mice (50/sex/dose) were administered doses of 0, 25, or 50 mg/kg-day (males) or 0, 75,
or 150 mg/kg-day (females) for 5 days/week for 102 weeks. All animals were subjected to gross
and microscopic examinations for neoplastic lesions. Survival of all dosed animals was
comparable to or greater than the corresponding control group. Statistically significant increases
were observed in the incidences of neoplasms of the large intestine and kidney in male and female
rats, the kidney in male mice, and the liver in female mice (Table VIII-6). The authors noted that
neoplasms of the large intestine and kidney are uncommon tumors in F344/N rats and B6C3F,
mice. They concluded that under the conditions of these 2-year gavage studies, clear evidence of
carcinogenic activity existed for male and female rats and mice.
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VIII-31
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Table VIII-6 Tumor Frequencies in Rats and Mice Exposed to Bromodichloromethane in
Corn Oil for 2 Years - Adapted from NTP (1987)
Animal
Tissue/Tumor
Tumor Frequency
Male rat
Control
50 mg/kg
100 mg/kg
Large
intestine3
Adenomatous polyp
0/50
3/49
33/50b
Adenocarcinoma
0/50
11/4 9b
38/50b
Combined
0/50
13/49b
45/50b
Kidneya
Tubular cell adenoma
0/50
1/49
3/50
Tubular cell adenocarcinoma
0/50
0/49
10/50b
Combined
0/50
1/49
13/50b
Female rat
Control
50 mg/kg
100 mg/kg
Large
intestinec
Adenomatous polyp
0/46
0/50
7/47b
Adenocarcinoma
0/46
0/50
6/47b
Combined
0/46
0/50
12/47b
Kidney
Tubular cell adenoma
0/50
1/50
6/5 0b
Tubular cell adenocarcinoma
0/50
0/50
9/5 0b
Combined
0/50
1/50
15/50b
Male mouse
Control
25 mg/kg
50 mg/kg
Kidneyd
Tubular cell adenoma
1/46
2/49
6/50
Tubular cell adenocarcinoma
0/46
0/49
4/50
Combined
1/46
2/49
9/5 0b
Female mouse
Control
75 mg/kg
150 mg/kg
Liver
Hepatocellular adenoma
1/50
13/48b
23/50b
Hepatocellular carcinoma
2/50
5/48
10/50b
Combined
3/50
18/48b
29/50b
a One rat died at week 33 in the low-dose group and was eliminated from the cancer risk calculation.
b Statistically significant at p<0.05, compared to controls.
c Intestine not examined in four rats from control group and three rats from high-dose group.
d In the control group, two mice died during the first week, one mouse died during week, nine and one escaped in week 79.
One mouse in the low-dose group died in the first week. All of these mice were eliminated from the cancer risk calculations.
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VIII - 32
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Use of the NTP rodent studies (NTP,1987) for derivation of cancer risk estimates for
bromodichloromethane is complicated by the use of a corn oil as a dosing vehicle. Although a
vehicle effect has not been reported for brominated trihalomethanes, it can be inferred from
studies of chloroform carcinogenicity that such an effect might exist, at least for hepatic tumors in
mice. Therefore, in the case of bromodichloromethane, the U.S. EPA believes that the most
appropriate basis of the cancer risk estimate is the incidence of renal tumors in male mice. Renal
tumors are considered to be appropriate because these tumors were increased in a dose-dependent
manner in both mice (male) and rats (both sexes).
c. Choice of Approach and Rationale
The LMS model (U.S. EPA, 1986) and the default linear approach described by the
Proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996; 1999) were used to
quantify the cancer risk associated with exposure to bromodichloromethane. Although data are
mixed, the weight of evidence indicates that bromodichloromethane is mutagenic (see Section
V.F.I). Furthermore, recent studies (Melnick, et al. 1998; George et al., 2002) suggest that
induction of hepatic tumors occurs at doses of bromodichloromethane that have marginal or no
effect on hepatocyte labeling index, indicating that regenerative hyperplasia is not required for
tumor induction. Thus, use of a linear approach was considered appropriate for quantification of
cancer risk associated with exposure to bromodichloromethane.
d. Cancer Potency and Risk Estimates
The available estimates for cancer risk associated with bromodichloromethane are
summarized in Table VIII-7. U.S. EPA (1994b) recommended use of a cancer potency estimate
of 6.2 x 10"2 (mg/kg-day)"1 as reported in IRIS (1993a). This value was derived in accordance
with the 1986 Guidelines for Carcinogenic Risk Assessment (U.S. EPA, 1986), based on the
occurrence of renal tumors in male mice. A unit risk of 1.8 x 10"6 was estimated using an
assumed body weight of 70 kg and a drinking water ingestion rate of 2 L. This estimate was used
to calculate a drinking water concentration of approximately 6 |ig/L associated with a 10"5 risk
(0.6 |ig/L for 10"6 risk).
A cancer potency value of 3.5 x 10"2 (mg/kg-day)"1 (U.S. EPA, 1998b) was derived using
the LMS method and an animal-to-human conversion factor of body weight374 (Table VIII-7). The
use of body weight34 is consistent with recommendations in U.S. EPA (1992b). This potency
factor is also based on the occurrence of renal tumors in male mice. A unit risk of 1 x 10"6 (|ig/L)"
1 was estimated for bromodichloromethane using an assumed body weight of 70 kg and a drinking
water ingestion rate of 2 L. This estimate was used to calculate a drinking water concentration of
approximately 10 |ig/L associated with a 10"5 risk (1 |ig/L for 10"6 risk).
Cancer risk estimates were also obtained using the LED10 (the lower 95% confidence limit
on a dose associated with 10% extra risk) for renal tumors in mice and assuming a linear mode of
action for the carcinogenicity of bromodichloromethane (Table VIII-7). A cancer potency value
of 3.4 x 10"2 (mg/kg-day)"1 was derived using this approach. Aunit risk of 9.6 x 10"7 (|ig/L)"' was
estimated for bromodichloromethane using an assumed body weight of 70 kg and a drinking
water ingestion rate of 2 L. This estimate was used to calculate a drinking water concentration of
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approximately 10 |ig/L associated with a 10"5 risk (1 |ig/L for 10"6 risk). These values are closely
similar to corresponding values derived using the LMS approach with body weight scaling to the
3/4 power.
Table VIII-7 Summary of Cancer Risk Estimates for Bromodichloromethane
Method of Estimation
Tumor Site
Species
Sex
Slope Factor
(mg/kg-day)1
Unit Risk
(Hg/L )"'
LED1c
(fig/kg-
day)
105 Risk
Concentration
(Hg/L)
LMS Method Using
BWs/4Conversion
(U.S. EPA, 1998b)
Liver
Mouse
F
6.9xl0"2
2.0xl0"6
-
5
Kidney
Rat
M
F
5.5xl0"3
6.1xl0"3
1.6xl0"7
1.7xl0"7
-
64
57
Mouse
M
3.5xl0"2
l.OxlO"6
-
10
Large
intestine
Rat
M
F
1.7xl0"2
6.1xl0"3
4.9xl0"7
1.7xl0"7
-
20
57
LMS Method Using
BW2/sConversion
U.S. EPA (1994b)*
Liver
Mouse
F
1.3X10"1
3.7xl0"6
-
3
Kidney
Rat
M
F
8.7xl0"3
9.5xl0"3
2.5xl0"7
2.7xl0"7
-
40
37
Mouse
M
6.2xl0"2
1.8xl0"6
-
6
Large
intestine
Rat
M
F
2.5xl0"2
4.9xl0"3
7.1xl0"7
1.4xl0"7
-
14
72
LED10/Linear Method
Liver
Mouse
F
6.5xl0"2
1.9x10"'
1.5xl03
5
(U.S. EPA, 1998b)
Kidney
Rat
M
F
8.1xl0"3
8.8xl0"3
2.3xl0"7
2.5xl0"7
1.2x10"
l.lxlO4
43
40
Mouse
M
3.4 x 10"2
9.6 xlO"7
3.0 x10s
10
Large
intestine
Rat
M
F
2.8xl0"2
l.OxlO"2
8xl0"7
3xl0"7
3.5x10s
9.6x10s
12
34
* Based on information adapted from IRIS (1993a)
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B. Dibromochloromethane
1. Noncarcinogenic effects
a. One-day Health Advisory
Four candidate studies that investigated the acute oral toxicity of dibromochloromethane
were available. These studies are summarized in Table VIII-8 (below). Bowman et al. (1978)
administered dibromochloromethane by gavage to ICR Swiss mice at doses ranging from 500 to
4,000 mg/kg and found that sedation and anesthesia resulted at doses of 500 mg/kg or higher.
NTP (1985) conducted a preliminary range-finding study in which F344/N rats and B6C3F, mice
were administered dibromochloromethane by gavage at doses ranging from 160 to 2,500 mg/kg
and found that death may result from doses at 310 mg/kg or higher in mice or rats. More
recently, Miiller et al. (1997) investigated the cardiotoxic effects of acute oral
dibromochloromethane exposure in male Wistar rats. In this study, rats administered doses of 83
or 167 mg/kg exhibited transient changes in cardiovascular parameters, while rats administered
doses of 333 or 667 mg/kg exhibited persistent alterations in at least one of the cardiovascular
parameters that lasted throughout the postexposure observation period. Finally, Korz and
Gatterman (1997) investigated the behavioral toxicity of acute oral dibromochloromethane
exposure in male golden hamsters and observed only transient effects on the behavioral
parameters investigated.
These studies were not considered adequate for deriving the One-day HA, since evaluated
more sensitive endpoints such as histopathology. Therefore, the Ten-day HA for
dibromochloromethane (0.6 mg/L) calculated below is recommended for use as the One-day HA.
b. Ten-day Health Advisory
Studies considered for derivation of the Ten-day HA for dibromochloromethane are
summarized in Table VIII-9 below. The key studies in this group are those of Aida et al. (1992a),
Condie et al. (1983), and Melnick et al. (1998). These studies reported effects on sensitive
endpoints and had data suitable for BMD analysis. Use of the remaining studies was limited by a
variety of considerations, including lack of data suitable for BMD analysis (Chu et al., 1982a;
NTP, 1996; Coffin et al., 2000), toxicological relevance or difficulty in interpretation of the most
sensitive endpoint (NTP, 1985; Munson et al. 1982), and occurrence of effects only at the frank
toxicity level (NTP, 1985).
Melnick et al. (1998) administered dibromochloromethane to female B6C3F, mice by
gavage for 5 days/week for 3 weeks and identified a NOAEL of 100 mg/kg-day (duration-
adjusted NOAEL of 71 mg/kg-day) and a LOAEL of 192 mg/kg-day (duration-adjusted LOAEL
of 137 mg/kg-day) based on histologic changes in the liver (hepatocyte hydropic degeneration).
BMD analysis calculated BMD and BMDL10 values of 112 and 68 mg/kg-day, respectively.
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Table VIII-8 Summary of Candidate Studies for Derivation of the One-day HA for Dibromochloromethane
Reference
Species
Route
Exposure
Duration
Dose
(mg/kg-day)
Result
Bowman et al.
(1978)
Mouse
ICR
Swiss
Gavage
(aqueous)
Single dose
500 - 4000
Sedation; anesthesia; increased
mortality
NTP (1985)
Rat
F344/N
Mouse
B6C3Fj
Gavage
(corn oil)
Single dose
160 - 2500
Lethargy; death
Miiller et al.
(1997)
Rat
Wistar
Gavage
(olive oil)
Single dose
83 - 667
Transient changes in blood
pressure; effects on cardiac
muscle contractility
Korz and
Gatterman (1997)
Hamster
Gavage
(olive oil)
Single dose
50
Transient changes in post-
treatment behavior
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Table VIII-9 Summary of Candidate Studies for Derivation of the Ten-day HA for Dibromochloromethane
Reference
Species
n
Dose
Route
Exposure
Endpoints
NOAEL
LOAEL
BMD
BMDL1C
Sex
Duration
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
Aida et al.
Rat
7
Males
Feed
1 month
Body weight, clinical
18.3
56
29
6.7
(1992a)
Wistar
0
signs, serum
(liver
(Liver cell
M, F
18
biochemistry,
histopathology)
vacuolation in
56
hematology, histology
females)
173
14
5.5
Females
(Liver cell
0
vacuolation in
34
males)
101
332
Chu et al.
Rat
10
0
Drinking
28 days
Clinical signs, serum
68
..
Not modeled
..
(1982a)
SD
0.7
water
biochemistry,
M
8.5
68
histology
Condie et
Mouse
8-
0
Gavage
14 days
Serum enzymes, PAH
74
147
3.5
1.6
al.
CD-I
16
37
(oil)
uptake in vitro,
(elevated ALT,
(Renal mesangial
(1983)
M
74
histology
decreased PAH,
hypertrophy)
147
liver and kidney
histopathology)
11
6.9
(hepatic
cytoplasmic
vacuolation)
Melnick et
Mouse
10
0
Gavage
3 weeks
Body and liver
100 (marginal)
192
112*
68
al.
B6C3Fj
50/37
(oil)
(5 d/wk)
weights, serum
(liver
(hepatic hydropic
(1998)
F
100/71
chemistry, liver
histopathology)
degeneration)
192/137
histology
417/298
Munson et
Mouse
8-
0
Gavage
14 days
Body and organ
50
125
Not modeled
__
al.
CD-I
12
50
(aqueous)
weights, serum
(decreased
(1982)
M, F
125
chemistry,
humoral
250
hematology, immune
immunity)
function
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VIII - 37 February 20, 2003
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Table VIII-9 (cont.)
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day)
NTP (1985)
Rat
F344/N
M, F
5
0
60
125
250
500
1000
Gavage
(oil)
14 days
Body weight, clinical
signs, gross necropsy
250
500 (FEL)
(mortality,
lethargy, gross
pathology)
Not modeled
NTP (1985)
Mouse
B6C3Fj
M, F
5
0
30
60
125
250
500
Gavage
(oil)
14 days
Body weight, clinical
signs, gross necropsy
60
125
(stomach lesions)
143
218
54
(stomach nodules
- males)
77
(stomach nodules
- females)
NTP (1996)
Rat
F344/N
M,F
10
Males
0
4
12
28
Females
0
6-7
17-20
48-48
Drinking
water
29 days
Body weight, serum
chemistry,
hematology, gross
necropsy, histology,
sperm evaluation
28
Not modeled
Coffin et al.
(2000)
Mouse
B6C3Fj
F
10
0
100
300
Gavage
(oil)
11 days
Relative liver wt.;
liver histopathology;
labeling index
-
100
Not modeled
--
*BMD and BMDL10 calculated using duration-adjusted doses
Abbreviations :FEL, Frank Effect Level; SD, Sprague-Dawley
No data
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These values are considerably higher (approximately 5- to 10-fold) than BMD and BMDL10
values calculated for hepatic effects using data from Condie et al. (1983) or Aida et al. (1992a).
Condie et al. (1983) administered dibromochloromethane by gavage to male CD-I mice
for 14 days and identified a NOAEL of 74 mg/kg-day and a LOAEL of 148 mg/kg-day based on
minimal to moderate liver and kidney injury. Histologic changes in the liver included focal
inflammation and cytoplasmic vacuolization similar to that observed in the study by Aida et al.
(1992a). Effects in the kidney included minimal to slight epithelial hyperplasia at the high dose
and minimal to slight mesangial hypertrophy at all (non-control) doses. Data for cytoplasmic
vacuolization and renal mesangial hypertrophy were analyzed by BMD modeling. The lowest
BMD and BMDL10 (3.5 and 1.6 mg/kg-day, respectively) among all candidate studies were
identified for mesangial hypertrophy. However, the pattern of dose-response for this endpoint
(0/16, 4/5, 7/10, 7/10 at doses of 0, 37, 74, and 147 mg/kg-day, respectively) resulted in generally
poor curve fits (0.10.49) and a high degree of model-dependence (See
summary of modeling results in Appendix A). Thus, confidence in the reliability of the BMD for
renal effects was low. The BMD and BMDL10 values for hepatic cytoplasmic vacuolization were
higher (11 and 6.9 mg/kg-day, respectively). These results were based on incidences of 1/16, 3/5,
4/10, and 8/10 at doses of 0, 37, 74, and 147 mg/kg-day, respectively.
The study by Aida et al. (1992a) was selected as the basis for derivation of the Ten-day
HA. In this study, Wistar rats of both sexes were administered microencapsulated
dibromochloromethane in the diet for one month. The dose levels ranged from 18.3 to 173.3
mg/kg-day for males and from 34.0 to 332.5 mg/kg-day for females. A NOAEL of 18.3 mg/kg-
day and a LOAEL of 56.2 mg/kg-day were identified based on histologic changes (cell
vacuolization, swelling, and single cell necrosis) in the livers of male rats. BMD analysis of data
for hepatic cell vacuolization calculated BMD and BMDL10 values of 29 and 6.7 mg/kg-day in
females and 14 and 5.5 mg/kg-day in males, respectively. The BDML10 for hepatic cell
vacuolization in male rats was selected for calculation of the 10-day HA because it was
considered the lowest reliable value based on examination of the raw data and modeling results.
The incidence of this endpoint was 0/7, 1/7, 3/7, and 7/7 at doses of 0, 18, 56, and 173 mg/kg-
day, respectively.
Based on the BMDL10 calculated from the data of Aida et al. (1992a), the Ten-day HA is
derived as follows:
(5.5 mg/kg-day) (10 kg)
Ten-day HA
0.55 mg/L (rounded to 0.6 mg/L)
(100) (1 L/day)
5.5 mg/kg-day
BMDL10 based on hepatic cell vacuolization in rats fed
dibromochloromethane for one month
10 kg
Assumed body weight of a child
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100 = Composite uncertainty factor based on NAS/OW
guidelines; includes a factor of 10 for interspecies
variation and a factor of 10 for protection of sensitive
human populations
1 L/day = Assumed water consumption of a 10-kg child
The Ten-day HA was calculated using the conventional NOAEL/LOAEL approach for
comparison with the value obtained using the BMD approach. The Aida et al. (1992a) study
identified a NOAEL of 18.3 mg/kg-day based on the absence of hepatic effects in rats. Using this
value and and the assumptions described above, the Ten-day HA would be 1.8 mg/L (rounded to
2 mg/L).
c. Longer-term Health Advisory
Four candidate studies were considered for derivation of the Longer-term HA for
dibromochloromethane. These studies are summarized in Table VIII-10 (below). Selected data
from three of these studies were modeled using the BMD approach. The results of BMD analysis
are included in Table VIII-11.
Chu et al. (1982b) administered dibromochloromethane to Sprague-Dawley rats in the
drinking water at doses ranging from 0.57 to 236 mg/kg-day. This study identified a NOAEL of
49 mg/kg-day, and a LOAEL of 224 mg/kg-day based on mild hepatic lesions (increased
cytoplasmic volume and vacuolation due to fatty infiltration) observed in males. BMD and
BMDL10 values of 18 and 5.3 mg/kg-day, respectively, were calculated for hepatic vacuolization
using the BMDS software. Daniel et al. (1990) identified a LOAEL of 50 mg/kg-day based on
hepatic lesions (centrilobular lipidosis) observed in male Sprague-Dawley rats and on kidney
lesions (tubular degeneration) observed in female Sprague-Dawley rats administered
dibromochloromethane by gavage for 90 consecutive days. BMD and BMDL10 values of 20 and
4.2 mg/kg-day, respectively, were calculated for renal tubular degeneration using the Crump
BMD software (K. S. Crump, Inc.). NTP (1985) administered doses of dibromochloromethane
ranging from 15 to 250 mg/kg-day to male and female mice. The compound was administered by
gavage in corn oil, five days per week for 13 weeks. NOAEL and LOAEL values of 125 and 250
mg/kg-day were identified on the basis of renal and hepatic lesions. BMD and BMDL10 values
were not calculated since lesions occurred only at the high dose of 250 mg/kg-day.
The NTP (1985) study conducted in rats was selected as the basis for derivation of the
Longer-term HA. In this study, F344/N rats were administered dibromochloromethane by gavage
at dose levels ranging from 15 to 250 mg/kg for 5 days/week for 13 weeks. Severe lesions and
necrosis of the kidney, liver, and salivary glands were observed primarily at the high dose.
However, males exhibited a dose-dependent increase in the frequency of clear cytoplasmic
vacuoles indicative of fatty metamorphosis in the liver. This effect reached statistical
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Table VIII-10 Summary of Candidate Studies for Derivation of the Longer-term HA for Dibromochloromethane
Reference
Species
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day) *
BMDL10
(mg/kg-day) *
Chu et al.
(1982b)
Rat
SD
M, F
20
Males
0
0.57
6.1
49
224
Drinking
water
90 days
Body weight, serum
chemistry, histology
49
224
(decreased weight
gain, mild hepatic
lesions)
18
5.3
(Liver lesions in
males)
Females
0
0.64
6.9
55
236
Daniel et al.
(1990)
Rat
SD
M, F
10
0
50
100
200
Gavage
(oil)
90 days
Body weight, clinical
signs, serum
biochemistry, gross
necropsy, histology
-
50
(hepatic and renal
lesions)
20+
4.2+
(kidney cortex
degeneration in
females)
NTP (1985)
Rat
F344/N
M, F
10
0
15
30
60
125
250
Gavage
(oil)
13 weeks
(5 d/wk)
Body weight, clinical
signs, histology
30
60
(hepatic lesions)
2.5
1.7
(liver fatty
metamorphosis in
males)
NTP (1985)
Mouse
B6C3Fj
M, F
10
0
15
30
60
125
250
Gavage
(oil)
13 weeks
(5 d/wk)
Body weight, clinical
signs, histology
125
250
(renal and hepatic
lesions)
Not modeled
* BMDL10 values were derived using duration-adjusted doses.
f Modeled using Crump benchmark dose software
Abbreviations: SD, Sprague-Dawley
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VIII-41 February 20, 2003
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significance at 60 mg/kg-day, and this dose was designated the LOAEL. The next lower dose (30
mg/kg-day) was designated as the NOAEL. BMD analysis using the BMDS program obtained
duration-adjusted BMD and BMDL10 values of 2.5 and 1.7 mg/kg-day, respectively. These
values were the lowest calculated among the three studies for which BMD analysis was
conducted.
Using the NTP (1985) rat study, the Longer-term HA for the 10-kg child is calculated as
follows:
(1.7 mg/kg-day)(10 kg)
Longer-term HA = =0.17 mg/L (rounded to 0.2 mg/L)
(100) (1 L/day)
where:
1.7 mg/kg-day = Duration-adjusted BMDL10 based on hepatic cell
vacuolization in rats exposed to dibromochloromethane by oil
gavage for 13 weeks
10 kg = Assumed body weight of a child
100 = Composite uncertainty factor based on NAS/OW guidelines;
includes a factor of 10 for interspecies variation and a factor
of 10 for protection of sensitive human populations
1 L/day = Assumed water consumption of a 10-kg child
The Longer-term HA for a 70-kg adult consuming 2 liters of water per day is calculated
according to the following equation:
(1.7 mg/kg-day)(70 kg)
Longer-term HA = =0.60 mg/L (rounded to 0.6 mg/L)
(100) (2 L/day)
For purposes of comparison, the Longer-term HAs were also calculated using the
conventional NOAEL/LOAEL approach. NTP (1985) identified a NOAEL of 30 mg/kg-day by
based on the absence of clinical signs or histologic alterations in rats exposed to
dibromochloromethane by corn oil gavage for 13 weeks. Using a duration adjusted dose of 21
mg/kg-day (obtained by multiplying the nominal dose by 5/7) and the assumptions for body
weight and drinking water ingestion described above, the Longer-term HAs would be 2.1 and 7.5
mg/kg-day for a 10 kg child and 70 kg adult, respectively.
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d. Reference Dose, Drinking Water Equivalent Level and Lifetime Health Advisory
This section reports the existing RfD value for dibromochloromethane and describes the
derivation of the RfD for this compound. This section also describes the calculation of Drinking
Water Equivalent Level and Lifetime Health Advisory values which require the RfD as input. For
this document, new and existing studies were reviewed and appropriate candidate data were
selected for benchmark dose (BMD) modeling. The results of BMD modeling were used in
conjunction with appropriate uncertainty factors to calculate the RfD. A comparison of the RfD
derived using the BMD approach to the results obtained using the conventional NOAEL/LOAEL
approach is also provided.
Description of the Existing RfD
The existing RfD for dibromochloromethane is 0.02 mg/kg-day (IRIS, 1992). This value
was derived using a duration-adjusted NOAEL of 21.4 mg/kg-day identified for the occurrence of
hepatic lesions in F344/N rats administered dibromochloromethane by corn oil gavage, 5
days/week for 13 weeks . An uncertainty factor of 1000 was used to account for extrapolation
from animal data, for protection of sensitive human subpopulations, and for use of a subchronic
study.
Identification of Candidate Studies for Derivation of the RfD
Candidate studies considered for derivation of the RfD for dibromochloromethane are
summarized in Table VIII-11 (below). Tobe et al. (1982) administered microencapsulated
dibromochloromethane in the diet to Wistar rats for 24 months at dose levels that ranged from 12
to 278 mg/kg-day. Although the study identified a NOAEL and a LOAEL of 12 and 49 mg/kg-
day, respectively, based on decreased body weight, changes in clinical chemistry parameters, and
gross liver appearance in males, a histopathological examination was not conducted. NTP (1985)
investigated the chronic oral toxicity of dibromochloromethane in F344/N rats and B6C3F, mice.
Only LOAEL values were identified in these studies. Specifically, the rat study identified a
LOAEL of 40 mg/kg-day based on histologic lesions in both male and female rats (e.g., fatty
change), and the mouse study identified a LOAEL of 50 mg/kg-day based on lesions in the liver
(fatty metamorphosis) and the thyroid (follicular cell hyperplasia) in the female mice. A thirteen-
week oral exposure study in rats (NTP, 1985) examined toxicity at a wider range of doses than
the chronic studies and identified NOAEL and LOAEL values of 30 and 60 mg/kg-day,
respectively, for histopathological changes in the liver.
Method of Analysis
Selected data from the candidate studies were analyzed using the benchmark dose (BMD)
modeling approach. Initially, data sets for potentially sensitive endpoints were selected as
described in U.S. EPA (1998b) and analyzed using the Crump Benchmark Dose Modeling
Software (K. S. Crump, Inc.). Results of this preliminary analysis are summarized in Table VIII-
12. Following the release of Version 1.2 of the BMDS program (U.S. EPA, 2000a), a subset of
the most sensitive endpoints identified using the Crump software was reanalyzed in accordance
with proposed U.S. EPA (2000b) recommendations. An advantage of analysis with the BMDS
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VIII - 43
February 20, 2003
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software is that several additional models are available to fit the data. The results of the analysis
using the BMDS software are included in Table VIII-11.
Choice of Principal Study and Critical Effect for the RfD
Two studies were identified as strong candidates for selection as the principal study. The
NTP (1985) subchronic study evaluated toxicological effects in male and female rats at five
concentrations of dibromochloromethane (15, 30, 60, 125, and 250 mg/kg-day) in addition to the
control. The chemical was administered by gavage in oil on five days per week for 13 weeks. A
relatively small sample size of 10 animals/treatment group was utilized. The endpoints evaluated
included clinical signs, body weight, serum biochemistry, and histopathological changes in organs.
NOAEL and LOAEL values of 30 and 60 mg/kg-day, respectively, were identified on the basis of
hepatic lesions using the conventional approach. Analysis of the data for fatty metamorphosis in
the liver using the BMDS program resulted in a duration-adjusted BMD of 2.7 mg/kg-day for this
endpoint, with a corresponding duration-adjusted BMDL10 of 1.7 mg/kg-day. A strength of this
study with respect to BMD modeling was the use of additional doses at the lower end of the
dose-response range. Inclusion of these doses permits more accurate characterization of the
shape of the dose-response curve and thus less uncertainty in the range of interest.
The second candidate for selection as the principle study for derivation of the RfD was the
chronic study conducted by NTP (1985). This study evaluated dibromochloromethane effects at
administered doses of 0, 40 and 80 mg/kg-day. The chemical was administered by gavage in oil
on five days per week for 104 weeks. The endpoints evaluated included clinical signs, body
weight, serum biochemistry, and histopathological changes in organs, and a LOAEL of 40 mg/kg-
day based on hepatic lesions was identified using the conventional approach. Analysis of the data
for fatty metamorphosis in the liver using the BMDS program resulted in a duration-adjusted
BMD of 2.5 mg/kg-day for this endpoint, with a corresponding duration-adjusted BMDL10 of 1.6
mg/kg-day based on duration adjusted doses.
A potential weakness of the NTP (1985) chronic study is the lack of dose-response
information at administered doses less than 40 mg/kg-day. A priori, the lack of information
regarding the shape of the curve at low doses would be expected to result in greater uncertainty
(and thus wider confidence limits) in the estimate of the chronic BMD. However, the BMD and
BMDL10 values calculated for fatty metamorphosis in the subchronic and chronic studies are
closely similar. This observation suggests that there is little potential for cumulative effects on the
occurrence of this lesion. The slight differences in the values may reflect both experimental
uncertainty and uncertainty in modeling.
Both studies were considered appropriate for derivation of the RfD. The NTP (1985)
investigation of chronic toxicity in rats was selected as the principle study on the basis of its
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February 20, 2003
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Table VIII-11 Summary of Candidate Studies for Derivation of the RfD for Dibromochloromethane
Reference
Species
Sex
n
Dose
(mg/kg-day)
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day) *
Tobe et al.
(1982)
Rat
Wistar
M, F
40
Males
0
12
49
196
Females
0
17
70
278
Diet
24 months
Body weight, serum
biochemistry, gross
pathology
12
49
(serum enzyme
changes and
altered liver
appearance)
Not modeled
NTP (1985)
Rat
F344/N
M, F
10
0
15
30
60
125
250
Gavage
(oil)
13 weeks
(5 d/wk)
Body weight, clinical
signs, serum
biochemistry, gross
necropsy, histology
30
60
(hepatic lesions)
2.5
1.7
(fatty
metamorphosis
in liver of
males)
NTP (1985)
Rat
F344/N
M, F
50
0
40
80
Gavage
(oil)
104 weeks
(5 d/wk)
Body weight, clinical
signs, gross necropsy,
histology
-
40
(hepatic lesions)
2.7
1.6
(fatty changes in
liver of males)
NTP (1985)
Mouse
B6C3Fj
M, F
50
0
50
100
Gavage
(oil)
105 weeks
(5 d/wk)
Body weight, clinical
signs, gross necropsy,
histology
50
(hepatic lesions)
9.V
7.1*
(thyroid
follicular cell
hyperplasia in
females)
Borzelleca
and
Carchman
(1982)**
Mouse
ICR
Swiss
10M
30F
0
17
171
685
Drinking
water
27 weeks
Maternal body
weight, gross
pathology, fetal
weight, survival,
teratogenicity
17
(marginal)
171
(maternal toxicity,
possible
fetotoxicity)
Not modeled
(insuff. data
provided in
publication)
--
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VIII - 45 February 20, 2003
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Table VIII-11 (cont.)
Reference
Species
Sex
n
Dose
(mg/kg-day)
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL1C
(mg/kg-day) *
NTP (1996)
**
Rat
SD
M, F
10
males
0
4.2
12.4
28.2
Group A
females
0
6.3
17.4
46.0
Group B
females
0
7.1
20.0
47.8
Drinking
water
29 days
Body weight, serum
chemistry,
hematology, gross
necropsy, histology,
sperm evaluation
28
Not modeled
* BMDL10 values were derived using duration-adjusted doses.
** These studies have been included in this table because they are reproductive/developmental studies and would be considered relevant for derivation of the
RfD. However, Borzelleca and Carchman (1982) found only marginal evidence for developmental toxicity at the low dose level and the NTP (1996)
study did not observe any reproductive or developmental effects at the dose levels evaluated.
f Modeled using Crump BMD software
Abbreviations: SD, Sprague-Dawley
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Table VIII-12 Results of Preliminary BMD Modeling of Selected Data from NTP (1985)
Studies
Study
Endpoint Modeled
BMDL10 (mg/kg-day) *
Subchronic NTP (1985) rat study
Fatty metamorphosis in liver of male rats
0.93
Chronic NTP (1985) rat study
Fatty metamorphosis in liver of male rats
1.16
Fatty metamorphosis in liver of female rats
No acceptable fit
"Ground glass" cytoplasm in liver of male
rats
4.93
Nephrosis in liver of female rats
17
Chronic NTP (1985) mouse study
Fatty metamorphosis in liver of female mice
7.68
Thyroid follicular cell hyperplasia in female
mice
7.09
* BMD modeling was conducted on duration-adjusted values using the Cramp BMD software.
longer duration. The critical endpoint is hepatotoxicity, as evidenced by the occurrence of fatty
metamorphosis in the livers of dibromochloromethane-treated animals. This effect was strongly
dose-dependent, with incidences of 27/50, 47/50, and 49/50 at the duration-adjusted doses of 0,
29, and 57 mg/kg-day. Selection of this study is strongly supported by the similar BMD
calculated for the same effect in the NTP (1985) subchronic study.
Derivation of the RfD
The duration-adjusted BMDL10 value from the chronic NTP (1985) rat study was selected
as the most appropriate basis for derivation of the RfD for dibromochloromethane. The RfD is
calculated using the following equation:
(1.6 mg/kg-day)
RfD = = 0.016 mg/kg-day (rounded to 0.02 mg/kg-day)
(100)
where:
1.6 mg/kg-day = Duration-adjusted BMDL10 based on fatty changes in the liver
of male rats
100 = Composite uncertainty factor based on NAS/OW guidelines;
includes a factor of 10 interspecies extrapolation and a factor
of 10 for protection of sensitive human populations
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A composite UF of 100 was used. The standard factors of 10 were used for interspecies
extrapolation and for protection of sensitive subpopulations. Furthermore, no additional
uncertainty factor was needed to account for an incomplete database. The database for
dibromochloromethane includes a two-generation study in ICR Swiss mice (Borzelleca and
Carchman 1982), a developmental toxicity study in Sprague-Dawley rats (Ruddick et al., 1983),
and a short-term reproductive and developmental toxicity study in rats (NTP, 1996). Therefore,
the database is considered nearly complete despite the lack of a developmental toxicity study in a
second species.
The DWEL for dibromochloromethane is calculated as follows:
(0.02 mg/kg-day) (70 kg)
DWEL = = 0.7 mg/L (700 |ig/L)
2 L/day
where:
0.02 mg/kg-day = RfD
70 kg = Assumed weight of an adult
2 L/day = Assumed water consumption by a 70-kg adult
Lifetime Health Advisory
The Lifetime Health Advisory (HA) represents that portion of an individual's total
exposure that is attributed to drinking water and is considered protective of noncarcinogenie
health effects over a lifetime of exposure. Dibromochloromethane is classified with respect to
carcinogenic potential as Group C: Possible human carcinogen. The Lifetime Health Advisory
(HA) is therefore calculated as follows:
(0.7 mg/L) (0.8)
Lifetime HA = = 0.06 mg/L (60 |ig/L)
10
where:
0.7 mg/kg-day = DWEL
0.8 = Relative Source Contribution (RSC), the
proportion of the total daily exposure
contributed by the dibromochloromethane in
drinking water
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10 = Uncertainty factor used in accordance with
U.S. EPA policy for Group C contaminants to
account for possible carcinogenicity
Alternative Approach for Derivation of the RfD
An alternative approach to the derivation of the RfD is use of the conventional
NOAEL/LOAEL method. The subchronic oral exposure study conducted by NTP (1985)
identified a NOAEL of 30 mg/kg-day. Using this value, a duration adjustment factor of 5/7, and a
composite uncertainty factor of 1000 (includes factors of 10 for interspecies extrapolation,
protection of sensitive subpopulations, and use of a subchronic study), the resulting RfD is 0.02
mg/kg-day (the same value as derived using the BMD approach). The corresponding DWEL is
0.7 mg/L, assuming an adult body weight of 70 kg and a drinking water ingestion rate of 2 L/day.
2. Carcinogenic Effects
a. Categorization of Carcinogenic Potential
Previous Evaluations of Carcinogenic Potential
The Carcinogenic Risk Assessment Verification Endeavor (CRAVE) group of the
U.S. EPA reviewed the available evidence on the carcinogenicity of the brominated
trihalomethanes and assigned dibromochloromethane to Group C: possible human carcinogen
(IRIS, 1992). This classification reflects inadequate human data and limited evidence of
carcinogenicity in animals.
Based on the 1996 Proposed Guidelines for Carcinogen Risk Assessment published in
1996 (U.S. EPA, 1996), dibromochloromethane is classified as cannot be determined. This
descriptor is considered appropriate when there are no or inadequate data in humans, and limited
evidence for carcinogenicity in animals.
IARC (1999c) has recently re-evaluated the carcinogenic potential of
dibromochloromethane. IARC concluded that there is limited evidence of carcinogenicity in
experimental animals and inadequate evidence in humans for dibromochloromethane.
Dibromochloromethane is therefore classified as Group 3: not classifiable as to carcinogenicity in
humans.
Categorization of Carcinogenic Potential Under the Proposed 1999 Cancer Guidelines
Cancer Hazard Summary
Under the proposed guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999)
dibromochloromethane shows suggestive evidence of carcinogenicity, but not sufficient to assess
human carcinogenic potential. This descriptor is appropriate when the evidence from human or
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animal data is suggestive of carcinogenicity, which raises a concern for carcinogenic effects but is
not judged sufficient for a conclusion as to human carcinogenic potential. This finding is based on
the weight of experimental evidence in animal models which indicate limited or equivocal
evidence of carcinogenicity.
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Supporting Information for Cancer Hazard Assessment
Human Data
The information on the carcinogenicity of dibromochloromethane from human studies is
inadequate. There are no epidemiological data specifically relating increased incidence of cancer
to exposure to dibromochloromethane. There are equivocal epidemiological data describing a
weak association of chlorinated drinking water exposures with increased incidences of bladder,
rectal, and colon cancer. U.S. EPA has determined that these studies cannot attribute the
observed effects to a single compound, as chlorinated water contains numerous other disinfection
byproducts that are potentially carcinogenic.
Animal Data
The carcinogenicity of dibromochloromethane in male and female animals has been
investigated in a well-designed and conducted corn oil gavage study conducted in rats and mice, a
dietary exposure study in rats, a drinking water study in mice, and a feed study in rats. No data
are available on the carcinogenic potential of dibromochloromethane administered via the
inhalation or dermal routes.
In the corn oil gavage study (NTP, 1985), the incidence of hepatocellular adenomas and
carcinomas and combined adenomas and carcinomas was significantly increased in high-dose
female mice and the incidence of hepatocellular adenomas was significantly increased in high dose
male mice. No evidence was observed for carcinogenicity in male or female rats under the
experimental conditions employed. Voronin (1987) did not observe significant increases in mice
treated with dibromochloromethane in drinking water for 104 weeks. Tobe et al. (1982) reported
no increase in gross tumors in rats treated exposed to dibromochloromethane in the diet for two
years.
Structural Analogue Data
Dibromochloromethane is structurally related to trihalomethanes that have shown varying
degrees of carcinogenic potential in rodents. Chloroform, the most extensively characterized
trihalomethane, is reported to be carcinogenic at high doses in several chronic animal bioassays,
with significant increases in the incidence of liver tumors in male and female mice and significant
increases in the incidence of kidney tumors in male rats and mice (U.S. EPA, 2001). The
occurrence of tumors in animals exposed to chloroform is demonstrably species-, strain-, and
gender-specific, and has only been observed under dose conditions that caused cytotoxicity and
regenerative cell proliferation in the target organ. The cancer database for structurally-related
brominated trihalomethanes is more limited, but includes well-conducted studies performed by the
National Toxicology Program. In a two-year corn oil gavage study of bromoform, NTP (1989a)
found clear evidence for carcinogenicity in female rats and some evidence of carcinogenicity based
on occurrence of tumors of the large intestine. In a two-year corn oil gavage study of
bromodichloromethane, NTP (1987) found clear evidence of carcinogenicity in male and female
rats (tumors of the large intestine), male mice (kidney tumors), and female mice (liver tumors). In
other bioassays, George et al. (2002) observed a significantly increased prevalence of neoplastic
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lesions in the liver of male rats at the lowest dose of bromodichloromethane administered in
drinking water, but not at higher doses. Tumasonis et al. (1985) reported significantly increased
incidences of hepatic neoplastic nodules, hepatic adenofibrosis, and lymphosarcoma in female rats
exposed to bromodichloromethane in drinking water.
Other Key Data
Dibromochloromethane is formed as a byproduct of drinking water disinfection with
chlorine. Exposure to dibromochloromethane may occur via ingestion of tap water, via dermal
contact during showering or bathing, or by inhalation of dibromochloromethane volatilized during
household activities. Absorption of single oral doses appears to be extensive.
Dibromochloromethane is rapidly metabolized and eliminated predominately as expired volatiles,
C02, or CO. Only a small amount (less than 10%) is eliminated in urine or in feces. No
comprehensive tissue data are available regarding the bioaccumulation or retention of
dibromochloromethane following repeated exposure. However, because of the rapid metabolism
and excretion of dibromochloromethane, marked accumulation and retention is not expected.
Dibromochloromethane itself is not directly reactive with DNA. Metabolism to reactive
species is a prerequisite for toxicity, as inferred from metabolic induction and inhibition studies.
In vitro and in vivo studies of the mutagenic and genotoxic potential of bromodichloromethane
have yielded both positive and negative results. Synthesis of the overall weight of evidence from
these studies is complicated by the use of a variety of testing protocols, different strains of test
organisms, different activating systems, different dose levels, different exposure methods (gas
versus liquid) and, in some cases, problems due to evaporation of the test chemical. Study results
for the mutagenicity of dibromochloromethane are mixed, and the overall evidence for
mutagenicity of this chemical is judged to be inconclusive (U.S. EPA, 1994b). Recent studies
conducted with strains of Salmonella engineered to express rat theta-class glutathione S-
transferase suggest that mutagenicity of the brominated trihalomethanes may be mediated by
glutathione conjugation.
Mode of Action
Limited or equivocal evidence has been obtained for the carcinogenic potential of
dibromochloromethane. Data to support primary mode of action for tumor development in the
liver of mice exposed to dibromochloromethane are lacking. In the absence of such information,
combined with an inconclusive weight-of-evidence evaluation for genotoxicity, the mode of action
for tumor development is assumed to be a linear process. The processes leading to tumor
formation in animals are expected to be relevant to humans.
Conclusion
Under the proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999)
dibromochloromethane shows suggestive evidence of carcinogenicity, but not sufficient to assess
human carcinogenic potential by the oral route. This weight-of-evidence evaluation is based on
1) limited or equivocal evidence of carcinogenicity in mice, but not rats, treated by oral pathways;
2) lack of epidemiological data specific to dibromochloromethane and equivocal data for drinking
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water drinking water exposures that cannot reliably be attributed to dibromochloromethane
among multiple other disinfection byproducts; 3) inconclusive results for many of the available
genotoxicity and mutagenicity tests; and 4) metabolism and mode of action that are reasonably
expected to be similar to those of structurally related compounds that induce tumors in
experimental animals. Although no cancer data exist for exposures via the dermal or inhalation
pathways, the weight-of-evidence conclusion is considered to be applicable to these pathways as
well. The finding for inhalation is based on the observation that patterns of metabolizing enzyme
activity in male rats for the related trihalomethane bromodichloromethane are similar for
exposure via the inhalation and gavage routes. Dibromochloromethane absorbed through the skin
is expected to be metabolized and cause toxicity in much the same way as dibromochloromethane
absorbed by the oral and inhalation routes.
b. Choice of Study for Quantification of Carcinogenic Risk
The proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999) do not
indicate dose-response assessment for chemicals for which there is suggestive evidence of
carcinogenicity, but not sufficient to assess human carcinogenic potential. However, the single
oral exposure study with positive tumor data for dibromochloromethane suggests significant
cancer potency for this compound in mice. A quantitative assessment of potency was therefore
considered appropriate.
In the absence of other carcinogenicity data, hepatic tumor incidence in female mice was
selected for estimation of carcinogenic risks associated with dibromochloromethane. These data
were obtained in an NTP (1985) study in which dibromochloromethane was administered in corn
oil to male and female B6C3F, mice (50 mice/sex/dose) by gavage 5 times/week for 104 to 105
weeks. The administered doses were 0, 50, or 100 mg/kg-day. Survival of dosed female mice
was comparable to that of the corresponding vehicle-control groups. High-dose male mice had
lower survival rates than the vehicle controls. At week 82, nine high-dose male mice died of an
unknown cause. An inadvertent overdose of dibromochloromethane given to low-dose male and
female mice at week 58 killed 35 male mice, but apparently did not affect the female mice. The
low-dose male mouse group was, therefore, considered to be unsuitable for analysis of neoplasms.
Compound-related nonneoplastic lesions were found in primarily in the livers of males
(hepatocytomegaly, necrosis, fatty metamorphosis) and females (calcification and fatty
metamorphosis). Nephrosis was also observed in male mice. Statistically significant increases in
the incidence of hepatocellular adenomas and in the combined incidence of adenomas and
carcinomas were observed in high-dose female mice. In male mice, a statistically significant
increase in the incidence of hepatocellular carcinomas and combined adenomas and carcinomas
was observed in the high-dose group; however, due to the overdose of dibromochloromethane in
the mid-dose group, the authors considered the tumor incidence data inadequate for tumor
analysis. Tumor incidence data from this study are presented in Table VIII-13.
c. Extrapolation model
The LMS model (U.S. EPA, 1986) and the default linear approach described by Proposed
Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996; 1999) were used to quantify the
risk associated with exposure to dibromochloromethane. Data for the mutagenicity and
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genotoxicity of dibromochloromethane are mixed (see Section V.F.2). U.S. EPA (1994b) has
previously determined that the weight of evidence for dibromochloromethane mutagenicity and
genotoxicity is inconclusive. At the present time there is insufficient evidence to establish with
certainty that dibromochloromethane exerts its carcinogenic effects via a non-genotoxic
mechanism. Thus, use of linear approaches was considered appropriate for quantification of
cancer risk associated with exposure to this compound.
Table VIII-13 Frequencies of Liver Tumors in Mice Administered Dibromochloromethane
in Corn Oil for 105 Weeks - Adapted from NTP (1985)
Treatment
Sex
Adenoma
Carcinoma
Adenoma or Carcinoma
(mg/kg-day)
(combined)
Vehicle Control
M
14/50
10/50
23/50
F
2/50
4/50
6/50
50
M
a
F
4/49
6/49
10/49
100
M
10/50
19/50b
27/50c
F
1 l/50b
8/50
19/50d
a Male low-dose group was inadequate for statistical analysis.
b p < 0.05 relative to controls.
c p < 0.01 (life table analysis); p = 0.065 (incidental tumor test) relative to controls.
d p < 0.01 relative to controls.
d. Cancer Potency and Unit Risk
The only tumor data available for dibromochloromethane are for liver tumors in female
B6C3F, mice (NTP, 1985). NAS (1987) previously utilized the tumor frequency data reported by
NTP (1985) to calculate an excess lifetime cancer unit risk of 8.3 x 10"7 The linearized
multistage model was utilized, with the assumption that 1 L of water per day containing 1 |ig/L of
dibromochloromethane was ingested. Based on this calculation, the concentration associated with
a risk of 10"6 is 0.6 |ig/L, assuming consumption of 2 L of water per day.
Other available estimates of cancer risks are summarized in Table VIII-14. U.S. EPA
(1994b) reported a slope factor of 8.4 x 10"2 (mg/kg-day)"1 calculated from the NTP (1985) data
in the absence of other appropriate tumorigenicity data for dibromochloromethane (IRIS, 1992).
This value was derived using the LMS model (extra risk) and a scaling factor of body weight273, as
specified in the 1986 Guidelines for Carcinogenic Risk Assessment (U.S. EPA, 1986). The
reported unit risk and 10"5 risk concentration were 2.4 x 10"6 (|ig/L)"' and 4 |ig/L, respectively.
A slope factor of 4.3 x 10"2 (mg/kg-day)"1 (U.S. EPA, 1998b) was derived using the LMS
model and a scaling factor of body weight3 4. The use of body weight3 4 as the scaling factor is
consistent with recommendations in (U.S. EPA., 1992b). A unit risk value of 1.2 x 10"6 (|ig/L)"'
was estimated using an assumed body weight of 70 kg and a drinking water ingestion rate of 2 L.
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This estimate was used to calculate a drinking water 10"5 risk concentration of 8 |ig/L (0.8 |ig/L at
10"6 risk).
Table VIII-14 Carcinogenic Risk Estimates for Dibromochloromethane
Method of
Estimation
Tumor
Site
Species
Sex
Slope Factor
(mg/kg-day)1
Unit Risk
(MS/I-)'
105 Risk
Cone. (ng/L)
LEDu
(Hg/kg-day)
LMS Method Using
BW3/4 Conversion
U.S. EPA (1998b)
Liver
Mouse
F
4.3xl0"2
1.2X10"6
8
-
LMS Method Using
BW2/3 Conversion
U.S. EPA (1994b)*
Liver
Mouse
F
8.4xl0"2
2.4X10"6
4
-
LED10/Linear Method
U.S. EPA (1998b)
Liver
Mouse
F
4.0xl0"2
1.2x10"'
9
2.5x10s
* Adapted from IRIS (1992)
Cancer risk estimates were also obtained using the LED10 (the lower 95% confidence limit
on a dose associated with 10% extra risk) for hepatic tumors and assuming a linear mode of
action for the carcinogenicity of dibromochloromethane (Table VIII-14). A cancer potency value
of 4.0 x 10"2 (mg/kg-day)"1 was derived using this approach. A unit risk of 1.2 x 10"6 (|ig/L)"' was
calculated using an assumed body weight of 70 kg and a drinking water ingestion rate of 2 L.
This estimate was used to calculate a drinking water concentration of 9 |ig/L associated with a 10"
5 risk (0.9 |ig/L for 10"6 risk). These values are similar to values derived using the LMS approach
with body weight scaling to the 3/4 power.
The use of a corn oil vehicle in the NTP (1985) study from which these data are derived
contributes uncertainty regarding the relevance of this value to exposure via drinking water. The
U.S. EPA plans to seek data on the tumorigenicity of dibromochloromethane in water in order to
clarify this issue.
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C. Bromoform
1. Noncarcinogenic effects
a. One-day Health Advisory
Acute toxicity information on bromoform is limited and no data suitable for BMD
modeling were identified. Some information is available on the former medicinal use of
bromoform in humans. In the past, oral doses of bromoform were used as a sedative for children
with whooping cough. Doses were typically one drop (approximately 180 mg) given three to six
times per day (Burton-Fanning, 1901). This treatment usually resulted in mild sedation in
children, although a few rare cases of death or near-death (believed to be due to accidental
overdoses) have been reported (e.g., Dwelle, 1903; Benson, 1907). Based on a dose of
540 mg/day given to a 10-kg child, the LOAEL for mild sedation is about 54 mg/kg-day.
Accordingly, the one day-HA for bromoform is calculated according to the following equation:
(54 mg/kg-day)(10 kg)
One-day HA = = 5.4 mg/L (rounded to 5 mg/L)
(100) (1 L/day)
where:
54 mg/kg-day = LOAEL based on sedation in children given oral doses of
bromoform
10 kg = Assumed weight of a child
100 = Composite uncertainty factor based on NAS/OW guidelines.
Includes a factor of 10 for interspecies variation and a factor
of 10 for protection of sensitive human populations
1 L/day = Assumed water consumption of a 10-kg child
b. Ten-day health Advisory
Candidate studies considered for derivation of the Ten-day HA are summarized in Table
VIII-15 (below). Condie et al. (1983) administered bromoform by gavage to male CD-I mice at
doses ranging from 72 to 289 mg/kg-day for 14 days and identified a NOAEL of 145 mg/kg-day
and a LOAEL of 289 mg/kg-day. The LOAEL is based on changes in clinical chemistry and on
minimal to moderate histologic changes in the kidney (intratubular mineralization, epithelial
hyperplasia, and mesangial hypertrophy and nephrosis) and in the liver (centrilobular pallor,
mitotic figures, focal inflammation, and cytoplasmic vacuolization). BMD modeling of data for
renal mesangial hypertrophy calculated BMD and BMDL10 values of 73 and 34 mg/kg-day,
respectively.
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Melnick et al. (1998) administered bromoform to female B6C3F, mice by gavage 5
days/week for 3 weeks and identified a NOAEL of 200 mg/kg-day (the lowest dose tested) and a
LOAEL of 500 mg/kg-day based on histologic changes in the liver (hepatocyte hydropic
degeneration). The duration-adjusted BMD and BMDL10 values for this endpoint were 146 and
104 mg/kg-day, respectively.
Munson et al. (1982) identified NOAEL and LOAEL values of 125 and 250 mg/kg-day,
respectively, based on elevated serum enzyme activity in mice. BMD modeling was not
conducted for this endpoint, since it was not considered a reliable basis for the Ten-day HA in the
absence of histopathological data. NTP (1989a) identified NOAEL and LOAEL values of 200
and 400 mg/kg-day, respectively, based on the occurrence of stomach nodules in rats and mice. A
BMD of 167 mg/kg-day was calculated for this endpoint in mice, with a corresponding BMDL10
of 66 mg/kg-day. However, occurrence of these nodules may represent a portal of entry effect.
Chu et al. (1982a) identified a freestanding NOAEL of 80 mg/kg-day in a drinking water study
conducted in rats. Coffin et al. (2000) identified a LOAEL of 200 mg/kg-day based on the
occurrence of liver histopathology and increased labeling index. The data of Coffin et al. were
not modeled because other studies used lower doses and were thus able to better characterize the
low-dose portion of the dose response curve.
The study conducted by Aida et al. (1992a) assessed toxicity in Wistar rats administered
bromoform microencapsulated in the diet at doses ranging from 56 to 728 mg/kg-day. The
duration of the study was one month. This study identified a NOAEL of 56 mg/kg-day and a
LOAEL of 208 mg/kg-day based on clinical chemistry changes and histologic changes in the liver
(cell vacuolization and swelling) of females. BMD modeling of results for liver cell vacuolization
in female rats calculated BMD and BMDL10 values of 16 and 2.3 mg/kg-day, respectively. These
were the lowest values observed among modeling results for candidate studies for the 10-day HA.
On this basis, and because histopathological changes in the liver are considered a sensitive
indicator of brominated trihalomethane toxicity, the study conducted by Aida et al. (1992a) was
considered the best choice for derivation of the Ten-day HA.
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Table VIII-15 Summary of Candidate Studies for Derivation of the Ten-day HA for Bromoform
Reference
Species
n
Route
Dose
Exposure
Endpoints
NOAEL
LOAEL
BMD
BMDL1C
Sex
Duration
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
(mg/kg-day)
Aida et al.
Rat
7
Feed
Males
1 month
Body weight,
62 males
187 males
Males
Males
(1992a)
Wistar
0
clinical signs,
56 females
208 females
140
51
M, F
62
serum
(serum chemistry
(Liver cell
187
biochemistry,
changes, liver
vacuolization)
618
hematology,
histopathology)
Females
histology
Females
Females
0
16
2.3
56
(Liver cell
208
vacuolization)
728
Chu et al.
Rat
10
Drinking
0.7
28 days
Clinical signs,
80
__
Not modeled
__
(1982a)
SD
water
8.5
serum
M
80
biochemistry,
histology
Condie et al.
Mouse
8-
Gavage
0
14 days
Serum enzymes,
145
289
73
34
(1983)
CD-I
16
(oil)
72
PAH uptake in
(elevated ALT,
(Renal mesangial
M
145
vitro, histology
decreased PAH,
nephrosis)
289
liver and kidney
histopathology)
Melnick et
Mouse
10
Gavage
0
3 weeks
Body and liver
200
500
146*
104*
al.
B6C3Fj
(oil)
200
(5 d/wk)
weights, serum
(liver
(Liver hydropic
(1998)
F
500
chemistry, liver
histopathology)
degeneration)
histology
Munson et
Mouse
6-
Gavage
0
14 days
Body and organ
125
250
Not modeled
__
al.
CD-I
12
(aqueous)
50
weights, serum
(elevated serum
(1982)
M, F
125
chemistry,
enzymes)
250
hematology,
immune function
NTP (1989a)
Mouse
5
Gavage
0
14 days
Body weight,
200
400
167
66
B6C3Fj
(oil)
50
clinical signs,
(stomach nodules)
(stomach
M
100
gross pathology
nodules)
200
400
600
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Table VIII-15 (cont.)
Reference
Species
Sex
n
Route
Dose
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day)
Coffin et al.
(2000)
Mouse
B6C3Fj
F
10
Gavage
(oil)
0
200
500
11 days
Relative liver
weight; liver
histopathology;
labeling index
-
200
(liver
histopathology;
labeling index)
Not modeled
-
NTP (1989a)
Rat
F344/N
M, F
5
Gavage
(oil)
0
100
200
400
600
800
14 days
Body weight,
clinical signs,
gross pathology
200
400
(decreased body
weight)
Not modeled
Ruddick et
al. (1983)**
Rat
SD
F
14-
15
Gavage
(oil)
0
50
100
200
Gestation
days 6-15
Body and organ
weights; maternal
serum chemistry;
hematology, and
histopathology;
developmental
parameters
50
100
(sternebral
aberrations)
50
33
(sternebral
aberrations)
* Duration-adjusted dose used to calculate BMD and BMDL10
** Ruddick et al (1983) is included because it is a reproductive study.
Abbreviations: SD, Sprague-Dawley
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Based on the BMDL10 identified in the Aida et al. (1992a) study, the Ten-day HA for a 10-
kg child is calculated according to the following equation:
(2.3 mg/kg-day)(10 kg)
Ten-day HA = =0.23 mg/L (rounded to 0.2 mg/L)
(100) (1 L/day)
where:
2.3 mg/kg-day = BMDL10 based on the occurrence of hepatic vacuolization in
female rats exposed to bromoform in the diet for one month
10 kg
= Assumed body weight of a child
100 = Uncertainty factor based on NAS/OW guidelines. Includes
a factor of 10 for interspecies variation and a factor of 10
for protection of sensitive human populations
1 L/day = Assumed water consumption of a 10-kg child
When the BMDL10 value for liver cell vacuolization in the Aida et al. (1992a) study is
used, the Ten-day HA for a 10-kg child is calculated to be 0.2 mg/L, assuming a drinking water
ingestion rate of 1 L/day and use of a composite uncertainty factor of 100. This value is slightly
lower than the Longer-term HA for a 10 kg child of 0.3 mg/L derived using subchronic data for
the same histopathological endpoint. This small difference may reflect experimental or BMD
modeling uncertainty.
For purposes of comparison, the Ten-day HA may also be derived using the conventional
NOAEL/LOAEL approach. The lowest LOAEL among the candidate studies was 100 mg/kg-
day for developmental effects in rats (Ruddick et al., 1983). The NOAEL in this study was 50
mg/kg-day. Aida et al. (1992a) identified NOAEL values of 56 and 62 mg/kg-day and LOAEL
values of 187 and 208 mg/kg-day for histopathological changes in male and female rats
administered bromoform in the diet. Chu et al. (1982a) identified a freestanding NOAEL of 80
mg/kg-day in rats. The data of Aida et al. (1992a) were selected for calculation of the Ten-day
HA because the study tested both male and female rats, incorporated more dose levels, and
identified both NOAEL and LOAEL values and because the NOAEL identified by Chu et al.
(1982a) is close to the lowest LOAEL of 100 mg/kg-day. Using the NOAEL of 62 mg/kg-day
for male rats and assuming the default body weight for a child (10 kg), the default drinking water
intake for a child (1 L/day), and a composite uncertainty factor of 100, the Ten-day HA would be
6.2 mg/L (rounded to 6 mg/L).
c. Longer-term Health Advisory
Candidate studies for derivation of the Longer-term HA are summarized in Table VIII-16
below. All studies identified histopathological changes in liver tissue as the critical toxicological
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Table VIII-16 Summary of Candidate Studies for Derivation of the Longer-term HA for Bromoform
Reference
Species
Sex
n
Route
Dose
Exposure
duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day) *
BMDL10
(mg/kg-day) *
Chu et al.
(1982b)
Rat
SD
M, F
20
Drinking
water
Male
0
0.65
6.1
57
218
90 days
Body weight, serum
chemistry, histology
57
218
(decreased weight
gain, mild hepatic
lesions)
Male
10
Females
No fit
Male
5.9
(Hepatic lesions)
Females
Females
0
0.64
6.9
55
283
NTP (1989a)
Rat
F344/N
M, F
10
Gavage
(corn oil)
0
12
25
50
100
200
13 weeks
(5 d/wk)
Body weight,
clinical signs, gross
necropsy, histology
25
50
(hepatic
vacuolization)
4.4
2.6
(hepatic
vacuolization in
male rats)
NTP (1989a)
Mouse
B6C3Fj
M, F
10
Gavage
(corn oil)
0
25
50
100
200
400
13 weeks
(5 d/wk)
Body weight,
clinical signs, gross
necropsy, histology
100
200
(hepatic
vacuolization)
88
55
(hepatic
vacuolization in
male mice)
Ruddick et
al. (1983)**
Rat
SD
F
9-
14
Gavage
(corn oil)
0
50
100
200
Gestation
days 6-15
Body and organ
weights; maternal
serum chemistry;
hematology, and
histopathology;
developmental
parameters
50
100
(sternebral
aberrations)
50
33
(sternebral
aberrations)
NTP (1989b)
Mouse
ICR
Swiss
M, F
20
Gavage
(corn oil)
0
50
100
200
105 days
Continuous breeding
reprod. study. Body
and organ weights,
histopathology,
reproductive
parameters
100
200
(decreased
maternal body
weight)
(not modeled)
* BMD and BMDL10 calculated using duration-adjusted doses
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** Ruddick et al (1983) is included because it is a reproductive study.
Not modeled
Abbreviations: SD, Sprague-Dawley
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effect. In one NTP (1989a) study, F344/N rats were administered bromoform by gavage at doses
ranging from 12 to 200 mg/kg-day for 5 days/week for 13 weeks. This study identified a NOAEL
of 25 mg/kg-day and a LOAEL of 50 mg/kg-day based on hepatic vacuolization observed in male
rats. BMD modeling with the BMDS program calculated a duration-adjusted BMD of 4.4
mg/kg-day (based on duration-adjusted doses), with a corresponding BMDL10 of 2.6 mg/kg-day.
These values were the lowest among the candidate studies.
In an analogous subchronic oral exposure study, NTP (1989a) exposed mice of both sexes
to doses of bromoform ranging from 25 to 400 mg/kg-day in addition to the control. This study
identified NOAEL and LOAEL values of 100 and 200 mg/kg-day, respectively, based on hepatic
vacuolization. BMD modeling with the BMDS program calculated a BMD of 88 mg/kg-day
(based on duration-adjusted doses), with a corresponding BMDL10 of 55 mg/kg-day. These
values were approximately 20-fold higher than the BMD and BMDL10 calculated for the NTP
(1989a) oral exposure study in rats.
Chu et al. (1982b) exposed rats of both sexes to bromoform in the drinking water for 90
days. The doses of bromoform ranged from 0.64 to 283 mg/kg-day in addition to the control.
This study identified NOAEL and LOAEL values of 57 and 218 mg/kg-day , respectively, based
on decreased weight gain and mild hepatic lesions in male mice. BMD modeling identified BMD
and BMDL10 values of 10 and 5.9 mg/kg-day using data for occurrence of hepatic lesions in male
mice. These values were approximately two-fold higher than the BMD and BMDL10 values
derived using data from the NTP (1989a) oral exposure study in rats. Strengths of this study
include exposure via drinking water, larger sample size (20 animals/treatment group), and the
administration of lower doses than used in the NTP (1989a) subchronic studies. Liver
histopathology data from this study were reported as combined lesions, with the types of lesions
described in the text.
The NTP (1989a) oral exposure study conducted in rats was selected for the derivation of
the Longer-term HA on the basis of the low values obtained for the BMD and BMDL10.
Selection of this study is strongly supported by the results of Chu et al. (1982b), which identified
slightly higher values in a drinking water study.
Based on the BMDL10 identified in the NTP (1989a) rat study, the Longer-term HA for a
10-kg child is calculated according to the following equation:
(2.6 mg/kg-day)(10 kg)
Longer-term HA = = 0.26 mg/L (rounded to 0.3 mg/L)
(100) (1 L/day)
where:
2.6 mg/kg-day = Duration-adjusted BMDL10 based on the occurrence of
hepatic vacuolization in male rats exposed to bromoform by
gavage for 13 weeks
10 kg = Assumed body weight of a child
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100 = Uncertainty factor based on NAS/OW guidelines; includes a
factor of 10 for interspecies variation and a factor of 10 for
protection of sensitive human populations
1 L/day = Assumed water consumption of a 10-kg child
The Longer-term HA for an adult consuming 2 liters of water per day is calculated
according to the following equation:
(2.6 mg/kg-day)(70kg)
Longer-term HA = = 0.91 mg/L (rounded to 0.9 mg/L)
(100) (2 L/day)
where:
2.6 mg/kg-day = Duration-adjusted BMDL10 based on the occurrence of
hepatic vacuolization in male rats exposed to bromoform by
gavage for 13 weeks
70 kg = Assumed body weight of an adult
100 = Composite uncertainty factor based on NAS/OW guidelines;
includes a factor of 10 for interspecies variation and a factor
of 10 for protection of sensitive human populations
2 L/day = Assumed water consumption of a 70-kg adult
For purposes of comparison, the Longer-term Health Advisories may also be derived
using the conventional NOAEL/LOAEL approach. The chronic oral exposure study conducted
by NTP (1989a) identified a NOAEL of 25 mg/kg-day based on the absence of clinical signs or
histological alterations in rats exposed to bromoform for 13 weeks. Using this value and
assuming default body weights (10 and 70 mg/kg-day, for adults and children, respectively),
default drinking water intake rates (1 and 10 L/day for children and adults, respectively), a
composite uncertainty factor of 100, and an exposure duration factor of 5/7, the Longer-term
HAs for a child and an adult would be 2 mg/L and 6 mg/L, respectively.
d. Reference Dose, Drinking Water Equivalent Level and Lifetime Health Advisory
This section reports the existing RfD value for bromoform and describes the derivation of
the RfD for this compound. This section also describes the calculation of Drinking Water
Equivalent Level and Lifetime Health Advisory values which require the RfD as input. For this
document, new and existing studies were reviewed and appropriate candidate data were selected
for benchmark (BMD) dose modeling. The results of BMD modeling were used in conjunction
with appropriate uncertainty factors to calculate the RfD. A comparison of the RfD derived using
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the BMD approach to the results obtained using the conventional NOAEL/LOAEL approach is
also provided.
Description of the Existing RfD
The existing RfD for bromoform is 0.02 mg/kg-day (IRIS, 1993b). This value was
derived using a duration-adjusted NOAEL of 17.9 mg/kg-day identified for the occurrence of
hepatic lesions in F344 rats administered dibromochloromethane by corn oil gavage 5 days/week
for 13 weeks (NTP, 1989a). An uncertainty factor of 1000 was used to account for extrapolation
from animal data, for protection of sensitive human subpopulations, and for use of a subchronic
study.
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Identification of Candidate Studies for the Derivation of the RfD
Three chronic exposure studies, a subchronic exposure study, a prenatal developmental
toxicity study, and a reproductive toxicity study were considered for derivation of the RfD for
bromoform. These studies are summarized in Table VIII-17 (below).
Tobe et al. (1982) administered bromoform microencapsulated in the diet to Wistar rats at
dose levels ranging from 22 to 619 mg/kg-day for 24 months. A NOAEL of 22 mg/kg-day and a
LOAEL of 90 mg/kg-day were identified based on gross liver lesions and changes in clinical
chemistry parameters in male rats.
NTP (1989a) conducted chronic oral exposure studies in rats and mice. In the rat study,
animals were administered bromoform by gavage in oil for 5 days/week for 103 weeks at doses of
100 or 200 mg/kg-day. This study identified the low dose as the LOAEL based on histologic
lesions in the liver (fatty change and chronic inflammation). In the mouse study, animals were
administered bromoform by gavage in corn oil, 5 days/week for 103 weeks at doses of 50 or 100
mg/kg-day for male mice and 100 or 200 mg/kg-day for female mice. Although no treatment-
related effects were observed in the male mice at the dose levels tested, treatment-related
histologic lesions in the liver (fatty changes) were observed in both low- and high-dose females.
Accordingly, this study identified a LOAEL of 100 mg/kg-day in female mice.
A subchronic oral exposure study conducted in rats (NTP, 1989a) was also considered as
a candidate for derivation of the RfD. This study utilized five doses of bromoform ranging from
12 to 200 mg/kg-day in addition to the control. The compound was administered to ten animals
per treatment group by gavage in corn oil, 5 days per week for 13 weeks. The endpoints
evaluated included body weight, clinical signs, gross necropsy, and histological changes. This
study identified a NOAEL and LOAEL of 25 and 50 mg/kg-day, respectively, on the basis of
histopathological changes (vacuolization) in the liver. The LOAEL identified in this study was the
lowest among all candidate studies.
The developmental study conducted by Ruddick et al. (1983) identified NOAEL and
LOAEL values of 50 and 100 mg/kg-day, respectively, for sternebral variations in the offspring of
female rats dosed with bromoform on gestation days 6-15. The reproductive toxicity study
reported by NTP (1989b) identified NOAEL and LOAEL values of 100 and 200 mg/kg-day for
reduced maternal body weight and decreased postnatal survival and liver histopathology in F,
mice of both sexes.
Method of Analysis
Selected data from the candidate studies were analyzed using the benchmark dose (BMD)
modeling approach. Initially, data sets for potentially sensitive endpoints were selected as
described in U.S. EPA (1998b) and analyzed using the Crump Benchmark Dose Modeling
Software (K. S. Crump, Inc.). Following the release of Version 1.2 of the BMDS program (U.S.
EPA, 2000a), data from the NTP (1989a) subchronic and chronic studies conducted in rats were
reanalyzed in accordance with proposed U.S. EPA (2000b) recommendations. An advantage of
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analysis with the BMDS software is that several additional models are available to fit the data.
The results of the analysis using the BMDS software are included in Table VIII-15.
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Table VIII-17 Summary of Candidate Studies for Derivation of the RfD for Bromoform
Reference
Species
Sex
n
Dose
Route
Exposure
Duration
Endpoints
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
BMD
(mg/kg-day)
BMDL10
(mg/kg-day) *
Tobe et al.
(1982)
Rat
Wistar
M, F
40
Male
0
22
90
364
Diet
24 months
Body weight, serum
chemistry, gross
pathology
22
90
(serum chemistry
changes, gross liver
lesions)
Not modeled
Female
0
38
152
619
NTP (1989a)
Rat
F344/N
M, F
10
0
12
25
50
100
200
Gavage
(corn oil)
13 weeks
(5 days/wk)
Body weight, clinical
signs, gross necropsy,
histology
25
50
(hepatic
vacuolization)
4.4
2.6
(Hepatic
vacuolization in
male rats)
NTP (1989a)
Rat
F344/N
M, F
50
0
100
200
Gavage
(corn oil)
103 weeks
(5 d/wk)
Body weight, clinical
signs, gross necropsy,
histology
100
(decreased body
weight, lethargy,
mild liver
histopathology)
13
1.4
(fatty changes in
liver of males)
NTP (1989a)
Mouse
B6C3Fj
M, F
50
Male
0
50
100
Gavage
(corn oil)
103 weeks
(5 d/wk)
Body weight, clinical
signs, gross necropsy,
histology
100 (male)
100 (female)
(decreased body
weight, mild liver
histopathology)
14.2t
10.6t
(fatty changes in
liver of females)
Female
0
100
200
Ruddick et al.
(1983)'
Rat
SD
F
9-
14
0
50
100
200
Gavage
(con oil)
Gestation
days 6-15
Developmental toxicity
study; body and organ
weights
50
(developmental)
200
(maternal)
100
(sternebral
variations)
50
33
(sternebral
variations)
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NTP (1989b)*
Mouse
50
0
Gavage
105 days
Continuous breeding
100
200
.
.
B6C3F!
50
(corn oil)
reprod. study. Body and
(decreased maternal
(not modeled)
M, F
100
200
organ weights,
histopathology,
reproductive parameters
body weight;
reduced postnatal
survival and liver
histopathology in ¥1
generation males
and females)
* BMD and BMDL10 values were derived using duration-adjusted doses.
f BMD modeled using the Crump BMD software (K. S. Crump, Inc.)
* These studies are included because they are reproductive/developmental studies.
Abbreviations: SD, Sprague-Dawley
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Choice of Principal Study and Critical Effect for the RfD
The subchronic study conducted by NTP (1989a) was selected for derivation of the RfD.
Two factors supported selection of this study. First, the critical effect identified in the subchronic
study was consistent with the critical effects identified in the chronic NTP studies (fatty changes in
liver of male rats and female mice). Second, BMD modeling of both data sets supported selection
of the subchronic study. BMD analysis using the BMDS program calculated a BMD of 13
mg/kg-day for fatty changes in the liver of males in the chronic study, with a corresponding
BMDL10 of 1.4 mg/kg-day. The lowest duration-adjusted dose in this study was 71 mg/kg-day
and the response at this dose was high (49/50). The magnitude of the difference between the
BMD and BMDL10 values thus reflects considerable uncertainty about the shape of the curve in
the low-dose region.
Duration-adjusted BMD and BMDL10 values for hepatic vacuolization in male rats of 4.4
and 2.6 mg/kg-day, respectively, were obtained using data for the subchronic study. This BMD is
approximately three-fold less than the BMD calculated from chronic data (above). The
availability of response data for three doses below 71 mg/kg-day (duration-adjusted dose) in the
subchronic study provided additional information about the shape of the dose-response curve in
the region of interest, and thus a more reliable estimate of the BMD. Although the BMDL10 value
for the subchronic study is higher than the value for the chronic study, this observation reflects
less uncertainty (smaller confidence interval) in the estimate of the subchronic BMD when the
results for the two studies are compared. The BMDL10 value calculated from the subchronic NTP
(1987) data was therefore selected for derivation of the RfD for bromoform.
The remaining studies were eliminated from consideration for the following reasons. The
study conducted by Tobe et al. (1982) did not identify a suitably sensitive endpoint
(histopathological examination was not conducted) and the data were never formally published or
submitted for peer review. The chronic study conducted by NTP (1989a) in mice reported mild
histopathological changes in female mice at a duration-adjusted dose of 71 mg/kg-day, the lowest
tested in this gender. However, both the BMD and BMDL10 were higher than those identified in
the subchronic study in rats, and these values were considered less reliable in the absence of data
at lower doses. The LOAELs identified in the developmental (Ruddick et al., 1983) and
reproductive (NTP, 1989a) studies were higher than the LOAEL values observed for
histopathology effects in the NTP subchronic study. The BMD and BMDL10 calculated for
sternebral variations in the Ruddick et al. (1983) were approximately 10-fold higher than those
identified in the subchronic study.
Derivation of the RfD
The duration-adjusted BMDL10 from the subchronic NTP (1989a) rat study was selected
for derivation of the RfD for bromoform. The RfD is calculated using the following equation:
(2.6 mg/kg-day)
RfD = = 0.03 mg/kg-day
100
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where:
2.6 mg/kg-day = Duration-adjusted BMDL10 based on hepatocellular
vacuolization in the liver of male rats
100 = Composite uncertainty factor based on NAS/OW
guidelines; includes a factor of 10 for interspecies
variation, a factor of 10 for protection of sensitive human
populations
A composite uncertainty factor of 100 was used in the calculation of the bromoform RfD.
The standard factors of 10 were used for interspecies extrapolation and for protection of sensitive
subpopulations. No uncertainty factor was added for extrapolation from a subchronic to a
chronic study because the BMD and BMDL10 for the subchronic study was either comparable to
or lower than the corresponding values from the chronic study. This observation suggests that a
cumulative effect on the liver does not occur for the endpoints examined. The database for
bromoform includes subchronic and chronic bioassays conducted in rats and mice (e.g. NTP
1989a), a two-generation reproductive toxicity study in mice (NTP 1989b), and a developmental
toxicity study in rats (Ruddick et al. 1983). Therefore, the database for bromoform was
considered sufficient and an uncertainty factor for database deficiencies was not included in the
calculation.
The DWEL for bromoform is calculated as follows:
(0.03 mg/kg-day) (70 kg)
DWEL = = 1.0 mg/L (1000 |ig/L)
2 L/day
where:
0.03 mg/kg-day = RfD
70 kg = assumed weight of an adult
2 L/day = assumed water consumption by a 70-kg adult
Lifetime Health Advisory
The Lifetime Health Advisory (HA) represents that portion of an individual's total
exposure that is attributed to drinking water and is considered protective of noncarcinogenie
health effects over a lifetime of exposure. Bromoform has been categorized with respect to
carcinogenic potential as Group B2: Probable human carcinogen (IRIS, 1993b). Therefore, in
accordance with U.S. EPA Policy, a Lifetime Health Advisory is not recommended.
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Alternative Approach for Derivation of the RfD
For comparison, the RfD can be calculated using the conventional NOAEL/LOAEL
approach. The subchronic NTP (1989a) oral exposure study identified a NOAEL of 25 mg/kg-
day based on absence of histopathological effects in rats exposed to dietary bromoform for 13
weeks. Using this value, a duration adjustment factor of 5/7, and an uncertainty factor of 1000
(including factors of 10 for interspecies extrapolation, protection of sensitive subpopulations, and
use of a subchronic study), the RfD would be 0.02 mg/kg-day. The corresponding DWEL would
be 0.7 mg/L assuming an adult body weight of 70 kg and a drinking water ingestion rate of 2
L/day.
2. Carcinogenic Effects
a. Categorization of Carcinogenic Potential
Previous Evaluations of Carcinogenic Potential
The Carcinogenic Risk Assessment Verification Endeavor (CRAVE) group of the
U.S. EPA has reviewed the available evidence on the carcinogenicity of bromoform and has
assigned it to Group B2: probable human carcinogen (IRIS, 1993b). Assignment to this category
is appropriate for chemicals where there are no or inadequate human data, but which have
sufficient animal data to indicate carcinogenic potential.
IARC (1999b) has recently re-evaluated the carcinogenic potential of bromoform. IARC
concluded that there is limited evidence of carcinogenicity in experimental animals and inadequate
evidence in humans for bromoform. Bromoform is therefore categorized as Group 3: not
classifiable as to carcinogenicity in humans.
Categorization of Carcinogenic Potential Under the Proposed 1999 Cancer Guidelines
Cancer Hazard Summary
Under the proposed guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999)
bromoform is likely to be carcinogenic to humans by all routes of exposure. This descriptor is
appropriate when the available tumor data and other key data are adequate to demonstrate
carcinogenic potential to humans. This finding is based on the weight of experimental evidence in
animal models which shows carcinogenicity by modes of action that are relevant to humans.
Supporting Information for Cancer Hazard Assessment
Human Data
The information on the carcinogenicity of bromoform from human studies is inadequate.
There are no epidemiological data specifically relating increased incidence of cancer to exposure
to bromoform. There are equivocal epidemiological data describing a weak association of
chlorinated drinking water exposures with increased incidences of bladder, rectal, and colon
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cancer. U.S. EPA has determined that these studies cannot attribute the observed effects to a
single compound, as chlorinated water contains numerous other disinfection byproducts that are
potentially carcinogenic.
Animal Data
The carcinogenicity of bromoform has been investigated in two species. These studies
include a well-designed and conducted corn oil gavage study conducted in rats and mice and a
study in which male Strain A mice were administered bromoform by intraperitoneal injection. No
data are available on the carcinogenic potential of bromoform administered via the inhalation or
dermal routes.
In the corn oil gavage study (NTP, 1989a), neoplasms of the large intestine (adenomatous
polyps or adenocarcinoma) were observed in male and female rats. The response for combined
adenoma and carcinoma reached statistical significance in female rats. The occurrence of tumors
of the large intestine in this study was considered biologically significant because they are
historically rare in rats. NTP (1989a) concluded that there was clear evidence for carcinogenicity
in females and some evidence of carcinogenicity in males. No evidence of bromoform
carcinogenicity was observed in male or female mice. Intraperitoneal injection of Strain A mice
with three concentrations of bromoform resulted in significantly increased tumor incidence only at
the middle dose tested.
Structural Analogue Data
Trihalomethanes structurally related to bromoform have shown varying degrees of
carcinogenic potential in rodents. Chloroform, the most extensively characterized trihalomethane,
is reported to be carcinogenic at high doses in several chronic animal bioassays, with significant
increases in the incidence of liver tumors in male and female mice and significant increases in the
incidence of kidney tumors in male rats and mice (U.S. EPA, 2001). The occurrence of tumors in
animals exposed to chloroform is demonstrably species-, strain-, and gender-specific, and has only
been observed under dose conditions that caused cytotoxicity and regenerative cell proliferation in
the target organ. The cancer database for structurally-related brominated trihalomethanes is more
limited, but includes well-conducted studies performed by the National Toxicology Program. In a
two-year corn oil gavage study of bromodichloromethane, NTP (1987) found clear evidence for
carcinogenicity in male and female rats (large intestine and kidney) and male (kidney) and female
(liver) mice. Tumasonis et al. (1987) reported a statistically significant increase in the incidence
of hepatic neoplastic nodules in rats exposed to bromodichloromethane in the drinking water. In
a two-year corn oil gavage study of dibromochloromethane, NTP (1985) determined that there
was some evidence of carcinogenicity in female mice and equivocal evidence of carcinogenicity in
male mice, based on the occurrence of hepatocellular adenomas and carcinomas. Other oral
exposure studies found no evidence for carcinogenicity of bromodichloromethane (Aida et al.,
1992b) or dibromochloromethane (Tobe et al., 1982; Voronin et al., 1987).
Other Key Data
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Bromoform is formed as a byproduct of drinking water disinfection with chlorine.
Exposure to bromoform may occur via ingestion of tap water, via dermal contact during
showering or bathing, or by inhalation of bromoform volatilized during household activities.
Absorption of single oral doses appears to be extensive. Bromoform is rapidly metabolized and
eliminated predominately as expired volatiles, C02, or CO. Only a small amount (less than 10%)
is eliminated in urine or in feces. No comprehensive tissue data are available regarding the
bioaccumulation or retention of bromoform following repeated exposure. However, because of
the rapid metabolism and excretion of bromoform, marked accumulation and retention is not
expected.
Bromoform itself is not directly reactive with DNA. Metabolism to reactive species is a
prerequisite for toxicity, as inferred from metabolic induction and inhibition studies. In vitro and
in vivo studies of the mutagenic and genotoxic potential of bromoform have yielded both positive
and negative results. Synthesis of the overall weight of evidence from these studies is complicated
by the use of a variety of testing protocols, different strains of test organisms, different activating
systems, different dose levels, different exposure methods (gas versus liquid) and, in some cases,
problems due to evaporation of the test chemical. However, because a majority of studies yielded
positive results, bromoform is considered to be at least weakly mutagenic and genotoxic. Recent
studies conducted with strains of Salmonella engineered to express rat theta-class glutathione S-
transferase suggest that mutagenicity of the brominated trihalomethanes may be mediated by
glutathione conjugation.
Mode of Action
The mode of action for tumor induction by bromoform has not been clearly elucidated and
may involve contributions from multiple bioactivation pathways. In each case, toxicity is believed
to result from interaction of reactive metabolites with cellular macromolecules. Proposed
bioactivation pathways for bromoform include: 1) production of reactive dihalocarbonyls by
oxidative metabolism; 2) production of reactive dihalomethyl radicals by oxidative metabolism;
and 3) formation of DNA-reactive species via a glutathione-dependent pathway. The relative
contribution of each pathway to tumor induction by bromoform has not been characterized. It is
possible that only the latter two processes lead to DNA damage in vivo, because the highly
reactive dihalocarbonyl intermediate may not survive long enough to enter the nucleus and react
with DNA. For this reason, cytotoxicity may be the primary consequence of the oxidative
pathway. Cytotoxicity coupled with regenerative hyperplasia is considered the primary mode of
action for tumor formation following exposure to high concentrations of chloroform, a
structurally-related trihalomethane which has low genotoxic potential. Data to support a similar
primary mode of action for tumor development in liver, kidney, and large intestine are currently
lacking for bromoform. In the absence of such information, combined with a positive weight-of-
evidence evaluation for genotoxicity, the mode of action for tumor development is assumed to be
a linear process. The processes leading to tumor formation in animals are expected to be relevant
to humans.
Conclusion
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Under the proposed guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999)
bromoform is likely to be carcinogenic to humans by the oral route. This weight-of-evidence
evaluation is based on 1) observations of tumors in rats treated by oral pathways; 2) lack of
epidemiological data specific to bromoform and equivocal data for drinking water drinking water
exposures that cannot reliably be attributed to bromoform among multiple other disinfection
byproducts; 3) positive results for a majority of the available genotoxicity and mutagenicity tests;
and 4) metabolism and mode of action that are reasonably expected to be comparable across
species. Although no cancer data exist for exposures via the dermal or inhalation pathways, the
weight-of-evidence conclusion is considered to be applicable to these pathways as well. The
finding for inhalation is based on the observation that patterns of metabolizing enzyme activity in
male rats are similar following exposure to a structurally-related compound
(bromodichloromethane) via the inhalation and gavage routes. Bromoform absorbed through the
skin is expected to be metabolized and cause toxicity in much the same way as bromoform
absorbed by the oral and inhalation routes.
b. Choice of Study for Quantification of Carcinogenic Risk
A single oral exposure study was available for the quantification of carcinogenic risk
associated with oral exposure to bromoform. NTP (1989a) conducted an oral exposure study in
B6C3F, mice and F344/N rats. No evidence of carcinogenicity was observed in male B6C3F,
mice exposed to bromoform via gavage (corn oil) at doses up to 100 mg/kg-day, or in female
mice exposed at doses up to 200 mg/kg-day for 5 days/week. Male and female F344/N rats (50
rats/sex/dose) were administered bromoform via gavage at doses of 0, 100, or 200 mg/kg-day for
5 days/week for 103 weeks (NTP 1989a). At termination, all animals were necropsied, and a
thorough histological examination of tissues was performed. Adenomatous polyps or
adenocarcinomas of the large intestine were noted in three high-dose male rats, eight high-dose
female rats, and one low-dose female rat (Table VIII-18). Despite the small number of tumors
found, the increase was considered biologically significant because these tumors are historically
rare in the rat. The study authors concluded that there was some evidence for carcinogenic
activity in male rats and clear evidence in female rats.
c. Extrapolation model
The LMS model (U.S. EPA, 1986) and the default linear approach described by the
Proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996; 1999) were used to
quantify the risk associated with exposure to bromoform. Although data are mixed , U.S. EPA
has previously concluded that the weight of evidence suggests that bromoform is mutagenic (see
Section V.F.3). At the present time, there are no data which indicate that bromoform-induced
tumorigenesis occurs as a consequence of cytotoxicity followed by regenerative hyperplasia.
Thus, use of a linear approach was considered appropriate for quantification of cancer risk
associated with exposure to bromodichloromethane.
Table VIII-18 Tumor Frequencies in Rats Exposed to Bromoform in Corn Oil for 2 Years -
Adapted from NTP (1989a)
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Animal
Tissue/Tumor
Tumor Frequency
Control
100 mg/kg
200 mg/kg
Male rat
Large intestine
Adenocarcinoma
0/50
0/50
1/50
Polyp (adenomatous)
0/50
0/50
2/50
Female rat
Large intestine
Adenocarcinoma
0/48
0/50
2/50
Polyp (adenomatous)
0/48
1/50
6/50
d. Cancer Potency and Unit Risk
Estimates of cancer risk associated with exposure to bromoform are summarized in Table
VIII-19. U.S. EPA (1994b) reported a cancer potency estimate of 7.9 x 10"3 (mg/kg-day)"1 for
bromoform based on the incidence of intestinal tumors in rats and derived using recommendations
in the 1986 Cancer Guidelines (U.S. EPA, 1986). The calculated value for unit risk is 2.3xl0"7
(|ig/L)"\ This estimate was used to calculate a drinking water concentration of 40 |ig/L
associated with a 10"5 risk.
A cancer potency estimate of 4.6xl0"3 (mg/kg-day)"1 (U.S. EPA, 1998b) based on the
incidence of intestinal tumors in rats was calculated using the LMS model and a scaling factor of
body weight34. Use of this scaling factor is consistent with recommendations in U.S. EPA
(1992b). Unit risk was estimated for bromoform using an assumed body weight of 70 kg and a
drinking water ingestion rate of 2 L. The calculated value for unit risk is 1.30x 10"7 (|ig/L)"\ This
estimate was used to calculate a drinking water concentration of 77|ig/L associated with a 10"5
risk (8 |ig/L for a risk of 10"6).
Cancer risk estimates were also obtained using the LED10 (the lower 95% confidence limit
on a dose associated with 10% extra risk) for hepatic tumors and assuming a linear mode of
action for the carcinogenicity of bromoform (Table VIII-19). A cancer potency value of 4.5 x 10"
3 (mg/kg-day)"1 was derived using this approach. A unit risk of 1.3 x 10"7 (|ig/L)"' was calculated
using an assumed body weight of 70 kg and a drinking water ingestion rate of 2 L. This estimate
was used to calculate a drinking water concentration of 78 |ig/L associated with a 10"5 risk (8
|ig/L for 10"6 risk). These values are similar to values derived using the LMS approach with body
weight scaling to the 3/4 power.
There are no data to suggest that tumor incidence in the large intestine is influenced by the
use of an oil vehicle. Therefore, the risk estimates reported above are believed to be applicable to
drinking water exposures.
Table VIII-19 Carcinogenic Risk Estimates for Bromoform
Method of
Tumor Site
Species
Sex
Slope Factor
Unit Risk
LEDjq
105 Risk
Concentration
(Hg/L)
Estimation
(mg/kg-day)1
0ig/L )"'
(Hg/kg-day)
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VIII - 76
February 20, 2003
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LMS Method Using
BW3'4 Conversion
U.S. EPA (1998b)
Large intestine
Rat
F
4.6xl0"3
1.3X10"7
-
77
U.S. EPA (1994b)*
Large intestine
Rat
F
7.9xl0"3
2.3xl0"7
-
40
LED10/Linear Method
U.S. EPA (1998b)
Large intestine
Rat
F
4.5xl0"3
1.3x10"'
2.2x10"
78
* Adapted from IRIS (1993b)
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D. Summary
Table VIII-20 Summary of Advisory Values for Bromodichloromethane,
Dibromochloromethane, and Bromoform
Advisory
Value
Reference
Bromodichloromethane
One-day HA for 10-kg child
1 mg/L
Narotsky et al. (1997)
Ten-day HA for 10-kg child
0.6 mg/L
NTP (1998)
Longer-term HA for 10-kg child
0.6 mg/L
CCC (2000d)
Longer-term HA for 70-kg adult
2 mg/L
CCC (2000d)
RfD
0.003 mg/kg-day
Aida et al. (1992b)
DWEL
100 ng/L
Aida et al. (1992b)
Lifetime HA
Not applicable
-
Oral Slope Factor c
3.5 x 10"2 (mg/kg-day)"1
NTP (1987)
Concentration for excess cancer risk (10"6)
1.0 ng/L
NTP (1987)
Unit Risk
1.0 x 10"6 (ng/L)"1
NTP (1987)
Dibromochloromethane
One-day HA for 10-kg child b
0.6 mg/L
Aida et al. (1992a)
Ten-day HA for 10-kg child
0.6 mg/L
Aida et al. (1992a)
Longer-term HA for 10-kg child
0.2 mg/L
NTP (1985)
Longer-term HA for 70-kg adult
0.6 mg/L
NTP (1985)
RfD
0.02 mg/kg-day
NTP (1985)
DWEL
700 ng/L
NTP (1985)
Lifetime HA
60 (ig/L
NTP (1985)
Oral Slope Factor c
4.3xlO"2 (mg/kg-day)"1
NTP (1985)
Concentration for Excess cancer risk (10"6)
0.8 ng/L
NTP (1985)
Unit Risk
1.2 x 10"6 (ng/L)"1
NTP (1985)
Bromoform
One-day HA for 10-kg child
5 mg/L
Burton-F arming (1901)
Ten-day HA for 10-kg child
0.2 mg/L
NTP (1989a)
Longer-term HA for 10-kg child a
0.2 mg/L
NTP (1989a)
Longer-term HA for 70-kg adult
0.9 mg/L
NTP (1989a)
RfD
0.03 mg/kg-day
NTP (1989a)
DWEL
1000 ng/L
NTP (1989a)
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Table VIII-20 (cont.)
Advisory
Value
Reference
Lifetime HA
Not applicable
--
Oral Slope Factor c
4.56xl0"3 (mg/kg-day)"1
NTP (1989a)
Concentration for Excess cancer risk (10"6)
8 ng/L
NTP (1989a)
Unit Risk
1.3 x 10"7 (ng/L)"1
NTP (1989a)
a The calculated value for the Longer-term HA was slightly higher than the values for the Ten-day HA.
Therefore, use of the Ten-day HA for a 10-kg child is recommended as an estimate of the Longer-term HA for a
10-kg child.
b Use of the Ten-day HA recommended as a conservative estimate of the One-day HA for a 10-kg child.
c Use of the Longer-term HA recommended as a conservative estimate of the Ten-day HA for a 10-kg child.
d The oral slope factor was calculated using the Linearized Multistage model and an animal-to-human scaling
factor of body weight3'4
Abbreviations: BW = Body weight; DWEL = Drinking water exposure limit; HA = Health advisory; LMS =
Linearized Multistage Model
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VIII - 79
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IX. REFERENCES
Abdel-Rahman, M.S. 1982. The presence of trihalomethanes in soft drinks. J. Appl. Toxicol.
2(3): 165-166. (As cited in U.S. EPA, 1994b.)
Aggazotti, G., G. Fantuzzi, E. Righi, and G. Predieri. 1998. Blood and breath analyses as
biological indicators of exposure to trihalomethanes in indoor swimming pools. Sci. Total
Environ. 217:155-163.
Ahmed, A.E., V.L. Kubic and M.W. Anders. 1977. Metabolism of haloforms to carbon
monoxide: I. In vitro studies. Drug Metab. Dispos. 5:198-204. (As cited in U.S. EPA, 1994b.)
Aida, Y., K. Takada, O. Uchida, K. Yasuhara, Y. Kurokawa and M. Tobe. 1992a. Toxicities of
microencapsulated tribromomethane, dibromochloromethane and bromodichloromethane
administered in the diet to Wistar rats for one month. J. Toxicol. Sci. 17(3): 119-133.
Aida, Y., K. Yasuhara, K. Takada, Y. Kurokawa and M. Tobe. 1992b. Chronic toxicity of
microencapsulated bromodichloromethane administered in the diet to Wistar rats. J. Toxicol. Sci.
17(2):51 -68, [Erratum] 17(3): 167.
Alavanja, M., I. Goldstein and M. Susser. 1978. A case-control study of gastrointestinal and
urinary tract cancer mortality and drinking water chlorination. In: Jolley R., H. Gorchev and
D.H. Hamilton Jr., eds. Water chlorination: Environmental impact and health effects. Vol. 2.
Ann Arbor, MI: Ann Arbor Science Publishers, pp. 395-409. (As cited in U.S. EPA, 1994b).
Allis, J.W., B.P. Anderson, G. Zhao, T.M. Ross, and R.A. Pegram. 2002. Evidence for the
involvement of CYP1A2 in the metabolism of bromodichloromethane in rat liver. Toxicology
176:25-37.
Allis, J.W., B.L. Brown, G. Zhao, and R.A. Pegram. 2001. The effects of inhalation exposure to
bromodichloromethane on specific rat CYP isoenzymes. Toxicology. 161:67-77.
Allis, J.W., and G. Zhao. 2002. Quantitative evaluation of bromodichloromethane metabolism by
recombinant rat and human cytochrome P450s. Chem. Biol. Interact. 140:137-153.
Anders, M.W., J.L. Stevens, R.W. Sprague, Z. Shaath and A.E. Ahmed. 1978. Metabolism of
haloforms to carbon monoxide: II. In vitro studies. Drug Metab. Dispos. 6:556-560.
Anderson, B.E., E. Zeiger, M.D. Shelby, M.A. Resnick, D.K. Gulati, J.L. Ivett and K.S. Loveday.
1990. Chromosome aberration and sister chromatid exchange test results with 42 chemicals.
Env. Mol. Mutagen. 16(Suppl. 18):55-137. (As cited in U.S. EPA, 1994b.)
Antoine, S.R., I.R. DeLeon and R.M. O'Dell-Smith. 1986. Environmentally significant volatile
organic pollutants in human blood. Bull. Environ. Contam. Toxicol. 36(3):364-371. (As cited in
U.S. EPA, 1994b.)
Draft -Do not cite or quote
IX- 1
February 20, 2002
-------
Armstrong, D.W. and T. Golden. 1986. Determination of distribution and concentration of
trihalomethanes in aquatic recreational and therapeutic facilities by electron-capture GC. LC-GC
Magazine of Liquid and Gas Chromatography 4(7):652-655. (As cited in U.S. EPA, 1994b.)
Aschengrau, A., S. Zierlier, and A. Cohen. 1989. Quality of community drinking water and the
occurrence of spontaneous abortions. Arch. Environ. Health. 44:283-90.
Aschengrau, A., S. Zierler, and A. Cohen. 1993. Quality of community drinking water and the
occurrence of late adverse pregnancy outcomes. Arch. Environ. Health. 48:105-113. (Abstract
only).
Ashley, D.L., Bonin, M.A., Cardinali, F.L, McCraw, J.M., and Wooten, J.V. (1994). Blood
concentrations of volatile organic compounds in a nonoccupationally exposed US population and
in groups with suspected exposure. Clin. Chem. 40:1401-1404.
Asoh, M., T. Tateishi, T. Kumai, and S. Kobayashi. 1998. Induction of CYP2B in foetal and
neonatal rats after maternal administration of phenobarbital. Pharmacol. Toxicol. 84:18-23.
Atlas, E. and S. Schauffler. 1991. Analysis of alkyl nitrates and selected halocarbons in the
ambient atmosphere using a charcoal preconcentration technique. Environ. Sci. Technol.
25(l):61-67. (As cited in U.S. EPA, 1994b.)
ATSDR. 1989. Agency for Toxic Substances and Disease Registry. Toxicological Profile for
Bromodichloromethane. PB/90/167461/AS. Available online at
www. atsdr. cdc. gov/toxprofiles/tp 129. pdf
ATSDR. 1990. Agency for Toxic Substances and Disease Registry. Toxicological Profile for
Bromoform and Chlorodibromomethane. PB/91/180323/AS. Available online at
www. atsdr. cdc. gov/toxprofiles/tp 13 0. pdf
Backer, L.C., D.L. Ashley, M.A. Bonin, F. Cardinali, S.M. Kieszak, and J.V. Wooten. 2000.
Household exposures to drinking water disinfection by-products: whole blood trihalomethane
levels. J. Expo. Anal. Environ. Epidemiol. 10:321-326.
Balster, R.L. and J.F. Borzelleca. 1982. Behavioral toxicity of trihalomethane contaminants of
drinking water in mice. Env. Health Perspect. 46:127-136. (As cited in U.S. EPA, 1994b.)
Barkely, J., J. Bunch, J.T. Bursey, N. Castillo, S.D. Cooper, J.M. Davis, M.D. Erickson, B.S.H.
Harris III, M. Kirkpatrick, L. Michael, S.P. Parks, E. Pellizzari, M. Ray, D. Smith, K. Tomer, R.
Wagner, and R.A. Zweidinger. 1980. Gas chromatography mass spectrometry computer analysis
of volatile halogenated hydrocarbons in man and his environment - a multimedia environmental
study. Biomed. Mass Spectrom. 7(4): 139-147. (As cited in Wallace, 1997).
Benson, M. 1907. Poisoning by bromoform. Br. Med. J. 204. (As cited in U.S. EPA, 1994b.)
Draft -Do not cite or quote
IX-2
February 20, 2002
-------
Berry, M. R., L.S. Johnson, K.P. Brenner, and K.W. Thomas. 1997. Dietary characterization in a
study of human exposures in the Lower Rio Grande Valley: II. Household waters. Environ. Intl.
23:693-703.
Bielmeier, S.R., D.S. Best, D.L. Guidici and M.G. Narotsky. 2001. Pregnancy loss in the rat
caused by bromodichloromethane. Toxicol. Sci. 59:309-315.
Boland, P. A. 1981. National screening program for organics in drinking water part II: Data.
SRI International. Prepared for U.S. Environmental Protection Agency. Contract No. 68-01-
4666. March. (As cited in U.S. EPA, 1994b.)
Borum, D. 1991. Telephone conversation (December 17) between Denis Borum, OST/OW,
U.S. Environmental Protection Agency, Washington, D.C., and Jeff Mosher, Wade Miller
Associates, Arlington, VA. (As cited in U.S. EPA, 1994b.)
Borzelleca, J.F. and R.A. Carchman. 1982. Effects of selected organic drinking water
contaminants on male reproduction. Publication no. EPA 600/1-82-009 (NTIS PB82-259847).
Research Triangle Park, NC: U.S. Environmental Protection Agency.
Bove, F. J. 1996. Public drinking water contamination and birthweight, prematurity, fetal deaths,
and birth defects. Toxicol. Ind. Health. 12(2): 255-66.
Bove F, Y., Shim, and P. Zeitz. 2002. Drinking water contaminants and adverse pregnancy
outcomes: A review. Environ. Health Perspect. 110 (Suppl l):61-74.
Bowman, F.J., J. Borzelleca and A.E. Munson. 1978. The toxicity of some halomethanes in
mice. Toxicol. Appl. Pharmacol. 44:213-215.
Brass, H.J. 1981. Rural water surveys organics data. Drinking Water Quality Assessment
Branch, Technical Support Division, Office of Drinking Water, U.S. Environmental Protection
Agency. Memo to Hugh Hanson, Science and Technology Branch, CSD, ODW, U.S.
Environmental Protection Agency. 1981. (As cited in U.S. EPA, 1994b.)
Brass, H.J., M.J. Weisner and B.A. Kingsley. 1981. Community water supply survey: Sampling
and analysis for purgeable organics and total organic carbon. Draft. American Water Works
Association Annual Meeting, Water Quality Division. June 9. (As cited in U.S. EPA, 1994b.)
Brenniman, G.R., J. Vasilomanolakis-Lagos, J. Amsel, T. Namekata and A.H. Wolff. 1980.
Case-control study of cancer deaths in Illinois communities served by chlorinated or
nonchlorinated water. In: Jolley, R. et al., eds. Water chlorination. Environmental impact and
health effects, Vol. 3. Ann Arbor, MI: Ann Arbor Science Publishers, pp. 1043-1057. (As cited
in U.S. EPA, 1994b.)
Brett, R.W. and R.A. Calverly. 1979. A one-year survey of trihalomethane concentration
changes within a distribution system. Am. Water Works Assoc. J. 71:515-520.
Draft -Do not cite or quote
IX-3
February 20, 2002
-------
Brodzinsky, R. and H.B. Singh. 1983. Volatile organic chemicals in the atmosphere: An
assessment of available data. U.S. Environmental Protection Agency, Office of Research and
Development, Research Triangle Park, NC.
Brzezinski, M.R., H. Boutelet-Bochan, R.E. Person, A.G. Fantel, and M.R. Juchau. Catalytic
activity and quantitation of cytochrome P-450 2E1 in prenatal human brain. J. Pharmacol. Exp.
Thera. 289:1648-1653. (Abstract only).
Bull, R.J., J.M. Brown, E.A. Meirerhenry, T.A. Jorgenson, M. Robinson and J.A. Stober. 1986.
Enhancement of the hepatotoxicity of chloroform in B6C3F1 mice by corn oil: Implications for
chloroform carcinogenesis. Env. Health Perspect. 69:49-58.
Bull, R.J. and F.C. Kopfler. Health effects of disinfectants and disinfectant by-products.
Prepared for the AWRA Research Foundation. August. (As cited in U.S. EPA, 1994b.)
Burton-Fanning, F.W. 1901. Poisoning by bromoform. Br. Med. J. 1202. (As cited in U.S.
EPA, 1994b.)
Cammann, K., and K. Huebner. (1995). Trihalomethane concentrations in swimmers' and bath
attendants' blood and urine after swimming or working in indoor swimming pools. Arch. Environ.
Health 50:61-65.
Cantor, K.P., R. Hoover, T.J. Mason and L.J. McCabe. 1978. Association of cancer mortality
with halomethanes in drinking water. J. Natl. Cancer Inst. 61:979-985. (As cited in U.S. EPA,
1994b.)
Cantor, K. P., C. F. Lynch, M. E. Hildesheim, M. Dosemeci, J. Lubin, M. Alavanja and G. Craun.
1998. Drinking water source and chlorination byproducts. I. Risk of bladder cancer.
Epidemiology 9(1): 21-8.
Carpenter, S.P., J.M. Lasker, and J.L. Raucy. 1996. Expression, induction, and catalytic activity
of the ethanol-inducible cytochrome P450 (CYP2E1) in human fetal liver and hepatocytes. Mol.
Pharmacol. 49(2):260-268.
CCC. 2000a. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories,
Raymond G. York, Study Director). Oral (Drinking Water) Range-finding Developmental
Toxicity Study of Bromodichloromethane (BDCM) in Rabbits. Protocol No. 2403-002P.
Arlington, VA: Chlorine Chemistry Council.
CCC. 2000b. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories,
Raymond G. York, Study Director). Oral (Drinking Water) Developmental Toxicity Study of
Bromodichloromethane in Rabbits. Protocol No. 2403-002. Arlington, VA: Chlorine Chemistry
Council.
CCC. 2000c. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories,
Raymond G. York, Study Director). Oral (Drinking Water) Range-finding Developmental
Draft -Do not cite or quote
IX-4
February 20, 2002
-------
Toxicity Study of Bromodichloromethane in Rats. Protocol No. 2403-001P. Arlington, VA:
Chlorine Chemistry Council.
CCC. 2000d. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories,
Raymond G. York, Study Director). Oral (Drinking Water) Developmental Toxicity Study of
Bromodichloromethane in Rats. Protocol No. 2403-003. Arlington, VA: Chlorine Chemistry
Council.
CCC. 2002. Chlorine Chemistry Council. (Prepared by Argus Research Laboratories, Raymond
G. York, Study Director). Oral (Drinking Water) Two-Generation (One-Litter per Generation)
Reproductive Study of Bromodichloromethane in Rats. Protocol No. 2403-001. Arlington, VA:
Chlorine Chemistry Council.
Chang, N.-W. and P.C. Singer. 1984. Formation and fate of bromoform during desalination. J.
Environ. Eng. 110(6): 1189-1193. (As cited in U.S. EPA, 1994b.)
Chen, W.J. and C.P. Weisel. 1998. Concentration changes of halogenated disinfection by-
products within a distribution system. Am. Water Works Assoc. J. 90:151-163.
Christian, M.S., R.G. York, A.M. Hoberman, R.M. Diener, and L.C. Fisher. 2001a. Oral
(drinking water) developmental toxicity studies of bromodichloromethane (BDCM) in rats and
rabbits. Int. J. Toxicol. 20:225-237.
Christian, M.S., R.G. York, A.M. Hoberman, R.M. Diener, L.C. Fisher, and G.A. Gates. 2001b.
Biodisposition of dibromoacetic acid and bromodichloromethane administered to rats and rabbits
in drinking water during range-finding reproduction and developmental toxicity studies. Int. J.
Toxicol. 20:239-253.
Christian, M.S., R.G. York, A.M. Hoberman, L.C. Fisher, and W.R. Brown. 2002. Oral
(drinking water) two-generation study of bromodichloromethane (BDCM) in rats. Int. J. Toxicol.
21:115-146.
Chu, I., V. Secours, I. Marino and D.C. Villeneuve. 1980. The acute toxicity of four
trihalomethanes in male and female rats. Toxicol. Appl. Pharmacol. 52:351-353.
Chu, I., D.C. Villeneuve, V.E. Secours and G.C. Becking. 1982a. Toxicity of trihalomethanes:
I. The acute and subacute toxicity of chloroform, bromodichloromethane, chlorodibromomethane
and bromoform in rats. J. Environ. Sci. Health. B17:205-224.
Chu, I., D.C. Villeneuve, V.E. Secours and G.C. Becking. 1982b. Trihalomethanes: II.
Reversibility of toxicological changes produced by chloroform, bromodichloromethane and
bromoform in rats. J. Environ. Sci. Health. B17:225-240.
Cocchioni, M., M.G. Pellegrini, L. Nacciarriti, and G. Bernacchia. 1996. Organo-halogenated
compounds in commercial prepared beverages. II. Igiene.Moderna. 106:287-293. [Italian],
Draft -Do not cite or quote
IX-5
February 20, 2002
-------
Coffin, J.C., R. Ge, S. Yang, P.M. Kramer, L. Tao, and M.A. Perieira. 2000. Effect of
trihalomethanes on cell proliferation and demethylation in female B6C3F1 mouse liver. Toxicol.
Sci. 58:243-252.
Condie, L.W., C.L. Smallwood and R.D. Laurie. 1983. Comparative renal and hepatotoxicity of
halomethanes: Bromodichloromethane, bromoform, chloroform, dibromochloromethane and
methylene chloride. Drug Chem. Tox. 6:563-578.
Cragle, D.L., C.M. Shy, R.J. Struba and E.J. Siff. 1985. A case-control study of colon cancer
and water chlorination in North Carolina. In: Jolley, R.L. et al., eds. Water chlorination.
Environmental impact and health effects, Vol. 5. Chelsea, MI: Lewis Publish. Inc., pp. 153-159.
(As cited in U.S. EPA, 1994b.)
Cresteil, T., P. Beaune, P. Kremers, C. Celies, F.P. Guengerich, and J.P. Leroux. 1985.
Immunoquantitation in of epoxide hydroxylase and cytochrome P450 isozymes in fetal and adult
liver microsomes. Eur. J. Biochem.l51:345-350. (As cited in Carpenter et al. 1996).
Daniel, F.B., M. Robinson, L.W. Condie and R.G. York. 1990. Ninety-day oral toxicity study of
dibromochloromethane in Sprague-Dawley rats. Drug Chem. Toxicol. 13(2-3): 135-154.
De Angelo, A.B., D.R. Geter, D.W. Rosenberg, C.K. Crary, and M.H. George. 2002. The
induction of aberrant crypt foci (ACF) in the colons of rats by trihalomethanes administered in
drinking water. Cancer Lett. 187:25-31.
De Groot, H. and T. Noll. 1989. Halomethane hepatotoxicity: Induction of lipid peroxidation
and inactivation of cytochrome P-450 in rat liver microsomes under low oxygen partial pressures.
Tox. Appl. Pharmacol. 97:530-537. (As cited in U.S. EPA, 1994b.)
DeMarini, D.M., M.L. Shelton, S.H. Warren, et al. 1997. Glutathione S-transferase-mediated
induction of GC—>AT transitions by halomethanes in Salmonella. Environ. Mol. Mutagen.
30(4):440-447.
Dodds, L. and W. King. 2001. Relation between trihalomethane compounds and birth defects.
Occup. Environ. Med. 58:443-446.
Dodds, L., W. King, C. Woolcott and J. Pole. 1999. Trihalomethanes in public water supplies
and adverse birth outcomes. Epidemiology. 10(3): 233-237.
Dwelle, E.H. 1903. Fatal bromoform poisoning. J.A.M.A. 41:1540. (As cited in U.S. EPA,
1994b.)
Entz, R.C., K.W. Thomas and G.W. Diachenko. 1982. Residues of volatile halocarbons in foods
using headspace gas chromatography. J. Agric. Food Chem. 30(5):846-849. (As cited in U.S.
EPA, 1994b.)
Draft -Do not cite or quote
IX-6
February 20, 2002
-------
Fair, P.S., R.C. Barth, J.J. Flesch and H.J. Brass. 1988. Measurement of disinfection
by-products in chlorinated drinking water. Proc. - Water Qual. Technol. Conf. 15:339-353. (As
cited in U.S. EPA, 1994b.)
French, A.S.,C.B. Copeland, D. Andrews, et al. 1999. Evaluation of the potential immunotoxicity
of bromodichloromethane in rats and mice. J. Toxicol. Environ. Health. 56(5):297-310.
Fujie, K., T. Aoki and M. Wada. 1990. Acute and subacute cytogenetic effects of the
trihalomethanes on rat bone marrow cells in vivo. Mutat. Res. 242(2): 111-119. (As cited in
U.S. EPA, 1994b.)
Fujie, K., T. Aoki, and S. Maeda. 1993. Sister chromatid exchanges induced by trihalomethanes
in rat erythroblastic cells and their suppression by crude catechin extracted from green tea.
Mutat. Res. 300(3-4):241-246.
Furlong, E.A.N, and F.M. D'itri. 1986. Trihalomethane: Levels in chlorinated Michigan drinking
water. Ecol. Model. June. 32(l-3):215-225. (As cited in U.S. EPA, 1994b.)
Gibori G, Khan I, Warshaw ML, McLean MP, Puryear TK, Nelson S, Durkee TJ, Azhar S,
Steinschneider A, Rao MC. Recent Progress in Hormone Research 44: 377-429, 1988.
Gallagher, M.D., J.R. Nuckols, L. Stallones, et al. 1998. Exposure to trihalomethanes and adverse
pregnancy outcomes. Epidemiology. 9(5):484-489.
Gao, P. and R.A. Pegram. 1992. In vitro hepatic microsomal lipid and protein binding by
metabolically activated bromodichloromethane. Toxicologist 12:421.
Gao, P., J.R. Thornton-Manning and R.A. Pegram. 1996. Protective effects of glutathione on
bromodichloromethane in vivo toxicity and in vitro macromolecular binding in Fischer 344 rats.
J. Toxicol. Environ. Health. 49:145-159.
Gargas, M.L., H.J. Clewell III, and M.E. Andersen. 1986. Metabolism of inhaled dihalomethanes
in vivo: Differentiation of kinetic constants for two independent pathways. Toxicol. Appl.
Pharmacol. 82:211-223.
Gebremichael, A., A.-M. Chang, A.R. Buckpitt, C.G. Plopper, and K.E. Pinkerton. 1995.
Postnatal development of cytochrome P4501A1 and 2B1 in rat lung and liver: Effect of aged and
diluted sidestream cigarette smoke. Toxicol. Appl. Pharmacol. 135:246-253.
George, M.H., G.R. Olson, D. Doerfler, T. Moore, S. Kilburn, and A. B. De Angelo. 2002.
Carcinogenicity of bromodichloromethane in drinking water to male F344/N rats and B6C3F1
mice. Int. J. Toxicol. 21:219-230.
Gottlieb, M.S., J.K. Carr, and D.T. Morris. 1981. Cancer and drinking water in Louisiana:
Colon and rectum. Int. Epidemiol. 10:117-125. (As cited in U.S. EPA, 1994b.)
Draft -Do not cite or quote
IX-7
February 20, 2002
-------
Guo Y, A.G. Hendrickx, J. Dieter, B.L. Lasley, D.R. Stewart, A.F. Tarantal, J.W. Overstreet.
1999. Biomarkers of early fetal loss in macaques: TCDD as a model of environmental toxicity,
Biology of Reproduction 60(3): 707-713.
Hakkola, J., E. Tanaka, and O. Pelkonen. 1998. Developmental expression of cytochrome
P450 enzymes in human liver. Pharmacol Toxicol 82 (5):209-217.
Hartwell, T.D., E.D. Pellizzari, R.L. Perritt, R.W. Whitmore, H.S. Zelon and L. Wallace. 1987.
Comparison of volatile organic levels between sites and seasons for the Total Exposure
Assessment Methodology (TEAM) study. Atmos. Environ. 21(11):2413-2424.
Hayashi, M., M. Kishi, T. Sofuni and M. Ishidate, Jr. 1988. Micronucleus tests in mice on 39
food additives and eight miscellaneous chemicals. Food Chem. Toxicol. 26(6):487-500. (As
cited in U.S. EPA, 1994b.)
Hayashi, M., H. Norppa, T. Sofuni, and M. Ishidate, Jr. 1992. Flow cytometric micronucleus
test with mouse peripheral erythrocytes. Mutagenesis 7(4)257-264.
Herren-Freund, S.L. and M.A. Pereira. 1986. Carcinogenicity of by-products of disinfection in
mouse and rat liver. Environ. Health Perspect. 69:59-66. (As cited in U.S. EPA, 1994b.)
Hewitt, W.R., E.M. Brown, and G.L. Plaa. 1983. Acetone-induced potentiation of
trihalomethane toxicity in male rats. Toxicol. Letters. 16:285-296.
Hildesheim, M. E., K. P. Cantor, C. F. Lynch, M. Dosemeci, J. Lubin, M. Alavanja and G. Craun
1998. Drinking water source and chlorination byproducts. II. Risk of colon and rectal cancers.
Epidemiology. 9(1): 29-35.
IARC. 1999a. International Agency for Research on Cancer. Bromodichloromethane. In:
IARC monographs on the evaluation of carcinogenic risks to humans: Re-evaluation of some
organic chemicals, hydrazine and hydrogen peroxide. Volume 71, Part Three. Lyon, France:
World Health Organization (WHO), International Agency for Research on Cancer, p. 1295-1304.
IARC. 1999b. International Agency for Research on Cancer. Bromoform. In: IARC
monographs on the evaluation of carcinogenic risks to humans: Re-evaluation of some organic
chemicals, hydrazine and hydrogen peroxide. Volume 71, Part Three. Lyon, France: World
Health Organization (WHO), International Agency for Research on Cancer, p. 1331-1338.
IARC. 1999c. International Agency for Research on Cancer. Chlorodibromomethane. In:
IARC monographs on the evaluation of carcinogenic risks to humans: Re-evaluation of some
organic chemicals, hydrazine and hydrogen peroxide. Volume 71, Part Three. Lyon, France:
World Health Organization (WHO), International Agency for Research on Cancer, p. 1309-1315.
Ioannides, C., C.R. Barnett, A. Irizar, et al. 1996. Expression of cytochrome P-450 proteins in
diseases. In: Ioannides, C., Ed. Cytochrome P-450. Metabolic and Toxicological aspects. Boca
Raton, FL: CRC Press, Inc., Chapter 12. pp.301-328
Draft -Do not cite or quote
IX-8
February 20, 2002
-------
Imaeda, K., Yamamoto, D., Watanabe, T., Yoshimura, Y., Uchiyama, K., Kato, K., and Ohsawa,
K. (1994). Formation of trihalomethanes during production of bean-curd. Jap. J. Toxicol.
Environ. Health 40:527-533.
IRIS. 1992. Integrated Risk Information System (December). Dibromochloromethane.
Washington, D.C.: U.S. Environmental Protection Agency.
IRIS. 1993a. Integrated Risk Information System. Bromodichloromethane. Washington, D.C.:
U.S. Environmental Protection Agency.
IRIS. 1993b. Integrated Risk Information System. Bromoform. Washington, D.C.: U.S.
Environmental Protection Agency.
Ishidate, M., T. Sofuni, K. Yoshikawa and M. Hayashi. 1982. Studies on the mutagenicity of
low boiling organohalogen compounds. Unpublished intraagency report to the National Institute
of Hygienic Sciences. Tokyo, Japan: Tokyo Medical and Dental University. (As cited in U.S.
EPA, 1994b.)
Jolley, R.L., G. Jones, W.W. Pitt and J.E. Thompson. 1978. Chlorination of organics in cooling
waters and process effluents. In: Jolley, R.L. et al., eds. Water chlorination. Environmental
impact and health effects. Vol. 1. Proceedings of the Conference on the Environmental Impact of
Water Chlorination. Ann Arbor, MI: Ann Arbor Science Publishers, Inc., pp. 105-138. (As cited
in U.S. EPA, 1994b.)
Jones, S. M., A.R. Boobis, G.E. Moore, and P. Stainer. 1992. Expression of CYP2E1 during
fetal development: methylation of the CYP2E1 gene in human fetal and adult liver samples.
Biochem. Pharmacol. 43:1876-1879. (As cited in Carpenter et al. 1996).
Jorgenson, T.A., E.F. Meierhenry, C.J. Rushbrook, R.J. Bull and M. Robinson. 1985.
Carcinogenicity of chloroform in drinking water to male Osborne-Mendel rats and female
B6C3F1 mice. Fund. Appl. Toxicol. 5:760-769.
Juchau, M.R., H. Boutelet-Bochan, and Y. Huang. 1998. Cytochrome-P450-dependent
biotransformation of xenobiotics in human and rodent embryonic tissues. Drug Metab. Rev.
30(3):541-568.
Juchau, M.R. and H.L. Yang. 1996. Cytochrome P450-dependent monooxygenation of
embryotoxic/teratogenic chemicals in human embryonic tissues. Fund. Appl. Toxicol. 34:166-
168. (As cited in Juchau et al. 1998).
Keegan, T.E., J.E. Simmons and R.A. Pegram. 1998. NOAEL and LOAEL determinations of
acute hepatotoxicity for chloroform and bromodichloromethane delivered in an aqueous vehicle to
F344 rats. J. Toxicol. Environ. Health 55(l):65-75.
Draft -Do not cite or quote
IX-9
February 20, 2002
-------
Kelley, R.D. 1985. Synthetic organic compound sampling survey of public water supplies. Iowa
Department of Natural Resources. Prepared for U.S. Environmental Protection Agency,
Washington, D.C., NTIS PB85-214427. (As cited in U.S. EPA, 1994b.)
Kerger, B.D., C.E. Schmidt, and D.J. Paustenbach. 2000. Assessment of airborne exposure to
trihalomethanes from tap water in residential showers and baths. Risk Anal. 20:637-651.
Komori, M., K. Nishio, T. Fujitani, H. Mitsukazu, K. Satoakimima, K. Itahashi, and T. Kamataki.
1989. Isolation of a new human fetal liver cytochrome P450 in human fetal liver cytochrome
P450 cDNA clone: evidence for expression of a limited number of forms of cytochrome P450 in
human fetal livers. Arch. Biochem. Biophys. 272:219-225. (As cited in Carpenter et al. 1996).
Kostiainen, R. (1995). Volatile organic compounds in the indoor air of normal and sick houses.
Atmos..Environ. 29:693-702.
King, W.D. and L.D. Marrett. 1996. Case-control study of bladder cancer and chlorination by-
products in treated water (Ontario, Canada). Cancer Causes Control. 7(6):596-604.
King, W.D., L. Dodds and A.C. Allen. 2000. Relation between stillbirth and specific chlorination
by-products in public water supplies. Environ. Health Perspect. 108:883-886.
Klinefelter, G.R., J.D. Suarez, N.L. Roberts and A.B. DeAngelo. 1995. Preliminary screening
for the potential of drinking water disinfection byproducts to alter male reproduction. Reprod.
Toxicol. 9(6):571-578.
Klotz, J.B. and L.A. Pyrch. 1999. Neural tube defects and drinking water disinfection by-
products. Epidemiology. 10(4): 3 83-3 90
Korz, V. and R. Gattermann. 1997. Behavioral alterations in male golden hamsters exposed to
chlorodibromomethane. Pharmacol. Biochem. Behav. 58(3):643-647.
Kramer, M.D., C.F. Lynch, P. Isacson and J.W. Hanson. 1992. The association of waterborne
chloroform with intrauterine growth retardation. Epidemiology. 3(5):407-413. (As cited in U.S.
EPA, 1994b.)
Krasner, S.W., M.J. McGuire, J.C. Jacangelo, N.L. Patania, K.M. Reagan and E.M. Aieta. 1989.
The occurrence of disinfection by-products in U.S. drinking water. Journal AWWA. August, pp.
41-53. (As cited in U.S. EPA, 1994b.)
Kroneld, R. and M. Reunanen. 1990. Determination of volatile pollutants in human and animal
milk by GC-MS. Bull. Environ. Contam. Toxicol. 44(6):917-923. (As cited in U.S. EPA,
1994b.)
Kurokawa, Y. 1987. Personal communication from Y. Kurokawa, National Institute of Hygienic
Sciences, Tokyo, Japan to R. Melnick, National Toxicology Program, North Carolina. (As cited
in IRIS, 1993b).
Draft -Do not cite or quote
IX- 10
February 20, 2002
-------
Landi, S. 2000. Mammalian class theta GST and susceptibility to carcinogens: a review. Mutat.
Res. 463(3):247-283.
Landi, S., N.M. Hanley, S.H. Warren, R.A. Pegram, and D.M. DeMarini. 1999. Induction of
genetic damage in human lymphocytes and mutations in Salmonella by trihalomethanes: role of
red blood cells and GSTT1-1 polymorphism. Mutagenesis. 14:479-482.
Larson, J.L., M.V. Templin, D.C. Wolf, K.C. Jamison, J. Leininger, S. Mery, K.T. Morgan, B.A.
Wong, R.B. Conolly, and B.E. Butterworth. 1996. A 90-day chloroform inhalation study in male
and female B6C3F, mice: Implications for cancer risk assessment. Fundam. Appl. Toxicol.
30:118-137.
Le Curieux, F., L. Gauthier, F. Erb, and D. Marzin. 1995. Use of the SOS chromotest, the Ames
fluctuation test, and the newt micronucleus test to study the genotoxicity of four trihalomethanes.
Mutagenesis. 10(4):333-341.
Lieber, L. 1997. Cytochrome P-4502E1: its physiological and pathological role. Physiol. Rev.
77(2):517-544.
Li X, D.C. Johnson, and K. K. Rozman. 1995. Effect of TCDD on estrous cyclicity and
ovulation in female Sprague-Dawley rats. Toxicol. Lett. 78: 219-222.
Lilly, P.D., J.E. Simmons and R.A. Pegram. 1994. Dose-dependent vehicle differences in the
acute toxicity of bromodichloromethane. Fundam. Appl. Toxicol. 23(1): 132-4.
Lilly, P.D., T.M. Ross, J.E. Simmons and R.A. Pegram. 1996. Effect of subchronic corn oil
gavage on the acute toxicity of orally administered bromodichloromethane. Toxicol. Letters. 87:
93-102.
Lilly, P.D., T.M. Ross and R.A. Pegram. 1997. Trihalomethane comparative toxicity: Acute
renal and hepatic toxicity of chloroform and bromodichloromethane following aqueous gavage.
Fundam. Appl. Toxicol. 40:101-110.
Lilly, P.D., M.E. Andersen, T.M. Ross, et al. 1998. A physiologically based pharmacokinetic
description of the oral uptake, tissue dosimetry, and rates of metabolism of
bromodichloromethane in the male rat. Toxicol. Appl. Pharmacol. 150(2):205-217.
Lindstrom, A. B., Pleil, J. D., and Berkoff, D. C. 1997. Alveolar breath sampling and analysis to
assess trihalomethane exposures during competitive swimming training. Environ. Health Perspect.
105:636-642.
Loveday, K.S., B.E. Anderson, M.A. Resnick and E. Zeiger. 1990. Chromosome aberration and
sister chromatid exchange tests in Chinese hamster ovary cells in vitro: V. Results with 46
chemicals. Environ. Mol. Mutagen. 16(4):272-303. (As cited in U.S. EPA, 1994b.)
Draft -Do not cite or quote
IX- 11
February 20, 2002
-------
Lynberg, M., J.R. Nuckols, P. Langois, D. Ashley, P. Singer, P. Mendola, C. Wilkes, H. Krapfl,
E. Miles, V. Speight, B. Lin, L. Small, A. Miles, M. Bonin, P. Zeitz, A. Tadkod, J. Henry, M.B.
Forrester. 2001. Assessing exposure to disinfection by-products in women of reproductive age
living in Corpus Cristi, Texas, and Cobb County, Georgia: descriptive results and methods. Env.
Health Perspect. 109(6):597-604.
Maddock, M.B. and J.J. Kelly. 1980. A sister chromatid exchange assay for detecting genetic
damage to marine fish exposed to mutagens and carcinogens. In: Jolley, R.H., W.A. Brungs and
R.B. Cumming, eds. Water chlorination: Environmental impact and health effects, Vol. 3. Ann
Arbor, MI: Ann Arbor Science, pp. 835-844. (As cited in U.S. EPA, 1994b.)
Magnus, P., J.J. Jaakkola, A. Skrondal, J. Alexander, G. Becher, T. Krogh, and E. Dybing. 1999.
Water chlorination and birth defects. Epidemiology. 10(5):513-517. (Abstract only).
Marcus, P. M., D. A. Savitz, R. C. Millikan and H. Morgenstern. 1998. Female breast cancer
and trihalomethane levels in drinking water in North Carolina. Epidemiology. 9(2): 156-60.
Mathews, J.M., P.S. Troxler and A.R. Jeffcoat. 1990. Metabolism and distribution of
bromodichloromethane in rats after single and multiple oral doses. J. Toxicol. Environ. Health.
30:15-22.
Matsuoka, A., K. Yamakage, H. Kusakabe, S. Wakuri, M. Asakura, T. Noguchi, T. Sugiyama, H.
Shimada, S. Nakayama, Y. Kasahara, Y. Takahashi, K.F. Miura, H. Hatanaka, M. Ishidate, Jr., T.
Morita, K. Watanabe, M. Hara, K. Odawara, N. Tanaka, M. Hayashi, and T. Sofuni. 1996. Re-
evaluation of chromosomal aberration induction on nine mouse lymphoma assay "unique positive"
NTP carcinogens. Mutat. Res. 369(3-4):243-52.
McGregor, D.B., A.G. Brown, S. Howgate, D. McBride, C. Riach, and W.J. Caspary. 1991.
Responses of the L5178Y mouse lymphoma cell forward assay: V.27 Coded chemicals. Environ.
Mol. Mutagen. 17(3): 196-219.
McGeehin, M.A., J.S. Reif, J.C. Becher, and E.J. Mangione. 1993. Case-control study of
bladder cancer and water disinfection methods in Colorado. Am. J. Epidemiol. 138(7):492-501.
McKone, T.E. 1987. Human Exposure to Volatile Organic Compounds in Household Tap Water:
The Indoor Inhalation Pathway. Environ. Sci. Technol. 27(12), 1194-1201.
McNeal, T.P., H.C. Hollifield, and G.W. Diachenko. 1995. Survey of trihalomethanes and other
volatile chemical contaminants in processed foods by purge-and-trap capillary gas
chromatography with mass selective detection. J. AOAC Int. 78:391-397.
Melnick, R.L., M.C. Kohn, J.K. Dunnick, et al. 1998. Regenerative hyperplasia is not required
for liver tumor induction in female B6C3F1 mice exposed to trihalomethanes. Toxicol. Appl.
Pharmacol. 148(1): 137-147.
Draft -Do not cite or quote
IX- 12
February 20, 2002
-------
Mera, N., S. Ohmori, Itahashi, S., Kiuchi, K., Igarashi, M, Rikihisa, T., Kitada, M. 1994.
Immunochemical evidence for the occurrence of Mu class glutathione S-transferase in human fetal
livers. J. Biochem (Tokyo) 116: 315-320.
Mills CJ, Bull RJ, Cantor KP, Reif J, Hrudey SE, Huston P (1998). Workshop report. Health
risks of drinking water chlorination by-products: Report of an expert working group. Chronic Dis
Can 19:91-102.
Mink, F.L., J. Brown and J. Rickabaugh. 1986. Absorption, distribution and excretion of 14C-
trihalomethanes in mice and rats. Bull. Environ. Contam. Toxicol. 37:752-758.
Mink, F.L., W.E. Coleman, J.W. Munch, W.H. Kaylor, and H.P. Ringhand. 1983. In vivo forms
of halogenated reaction products following peroral sodium hypochlorite. Bull. Environ. Contam.
Toxicol. 30:394-399. (As cited in U.S. EPA, 1984).
Miyagawa, M., H. Takasawa, A. Sugiyama, Y. Inoue, T. Murata, Y. Uno, and K. Yoshikawa.
1995. The in vivo-in vitro replicative DNA synthesis test (RDS) with hepatocytes prepared from
male B6C3F1 mice as an early prediction assay for putative nongenotoxic (Ames-negative) mouse
hepatocarcinogens. Mutat. Res. 343(2-3): 157-158.
Moon, S., J.M. Adamski and A. Lichtman. 1990. Profile of volatile organic chemicals in Nassau
County groundwater. In: Friedman, D. ed. Waste testing and quality assurance: Second volume.
ASTM Special Technical Publication, No. 1062. Philadelphia, PA: American Society for Testing
and Materials, pp. 50-63. (As cited in U.S. EPA, 1994b.)
Morimoto, K. and A. Koizumi. 1983. Trihalomethanes induce sister chromatid exchanges in
human lymphocytes in vitro and mouse bone marrow cells in vivo. Environ. Res. 32:72-79. (As
cited in U.S. EPA, 1994b.)
Morris, R.D., A.-M. Audet, I.F. Angelillo, T.C. Chalmers and F. Mosteller. 1992. Chlorination,
chlorination by-products, and cancer: A meta-analysis. Am. J. Public Health. 82(7):955-963.
(As cited in U.S. EPA, 1994b.)
Miiller, S.P., P. Wolna, U. Wiinscher and D. Pankow. 1997. Cardiotoxicity of
chlorodibromomethane and trichloromethane in rats and isolated rat cardiac myocytes. Arch.
Toxicol. 71:766-777.
Munson, A.E., L.E. Sain, V.M. Sanders, B.M. Kauffmann, K.L. White, Jr., D.G. Page, D.W.
Barnes and F. Borzelleca. 1982. Toxicology of organic drinking water contaminants:
Trichloromethane, bromodichloromethane, dibromochloromethane and tribromomethane.
Environ. Health Perspect. 46:117-126.
Nakajima, T., E. Elovaara, T. Okino, H.V. Gelboin, M. Klockars, V. Riihimaki, T. Aoyama, and
H. Vainio. 1995. Different cobtributions of cytochrome P450 2E1 and 2B1/2 to chloroform
hepatotoxicity in rat. Toxicol. Appl. Pharmacol. 133:215-222.
Draft -Do not cite or quote
IX- 13
February 20, 2002
-------
Narotsky, M.G., R.A. Pegram and R.J. Kavlock. 1997. Effect of dosing vehicle on the
developmental toxicity of bromodichloromethane and carbon tetrachloride in rats. Fundam.
Appl. Toxicol. 40:30-36.
Nelson, D.R., L. Kuymans, T. Kamataki, J.J. Stegeman, R. Feyereisen, D.J. Waxman, M.R.
Waterman, O. Gotoh, M.J. Coon, R.W. Estabrook, I.C. Gunsalus, and D,W. Nebert. 1996.
P450 superfamily: update on new sequences, gene mapping, accession numbers, and
nomenclature. Pharmacogenetics. 6:1-42.
Nestmann, E.R. and E.G.H. Lee. 1985. Genetic activity in Saccharomyces cerevisiae of
compounds found in effluents of pulp and papermills. Mutat. Res. 155:53-60. (As cited in U.S.
EPA, 1994b.)
NRC. 1999. National Research Council. Risk Assessment of Radon in Drinking Water.
Washington, D.C.: National Academy Press.
NTP. 1985. National Toxicology Program. Toxicology and carcinogenesis studies of
chlorodibromomethane in F344/N rats and B6C3F1 mice (gavage studies). Technical Report
Series No. 282. Research Triangle Park, NC: U.S. Department of Health and Human Services.
NTP. 1987. National Toxicology Program. Toxicity and carcinogenesis studies of
bromodichloromethane (CAS No. 75-27-4) in F344/N rats and B6C3F1 mice (gavage studies).
Technical Report Series No. 321. Research Triangle Park, NC: U.S. Department of Health and
Human Services.
NTP. 1989a. National Toxicology Program. Toxicology and carcinogenesis studies of
tribromomethane (bromoform) in F344/N rats and B6C3F1 mice (gavage studies). Technical
Report Series No. 350. Research Triangle Park, NC: U.S. Department of Health and Human
Services.
NTP. 1989b. National Toxicology Program. Bromoform: Reproduction and fertility assessment
in Swiss CD-I mice when administered by gavage. Report no. NTP-89-068. Research Triangle
Park, NC: National Institute of Environmental Health Sciences.
NTP. 1996. National Toxicology Program. Final report on the short term reproductive and
developmental toxicity of chlorodibromomethane (CAS No. 124-48-1) administered in drinking
water to Sprague-Dawley rats. Report by R.O.W. Sciences, Inc., Gaithersburg, MD to National
Toxicology Program. Publication no. NTIS/PB97-111728.
NTP. 1998. National Toxicology Program. Final Report on the short reproductive and
developmental toxicity of bromodichloromethane (CAS No. 75-27-4) administered in drinking
water to Sprague-Dawley rats. Research Triangle park, NC: National Institute of Environmental
Health Sciences. Publication no. NTIS/PB99-111262.
Draft -Do not cite or quote
IX- 14
February 20, 2002
-------
Nuckols, J.R., L. Stallones, and J. Reif. 1995. Evaluation of the use of a geographic information
sustem in drinking water epidemiology. In: Assessing and managing Risks from Drinking Water
Contamination: Approaches and Applications. IAHS Publ. 233, pp. 111-122.
Omiencinski, C.J., C. Hassett, and P. Costa. (1990). Developmental expression and in situ
localization of the phenobarbital-inducible rat hepatic mRNAs for cytochromes CYP2B1,
CYP2B2, CYP2C6, and CYP3A1. Mol. Pharmacol. 38(2):462-470.
Pankow, D., B. Damme, U. Wunscher, et al. 1997. Chlorodibromomethane metabolism to
bromide and carbon monoxide in rats. Arch. Toxicol. 71(4):203-210.
Pegram, R.A., M.E. Andersen, S.H. Warren, et al. 1997. Glutathione S-transferase-mediated
mutagenicity of trihalomethanes in Salmonella typhimurium: contrasting results with
bromodichloromethane off chloroform. Toxicol. Appl. Pharmacol. 144(1): 183-188.
Pellizzari, E.D. 1979. Organic screening in Lake Charles, LA, using gas chromatography mass
spectrometry computer techniques. EPA contract no. 68-02-2714. Research Triangle Institute,
Research Triangle Park, NC. (As cited in Brodzinsky and Singh, 1983).
Pellizzari, E.D. and J.E. Bunch. 1979. Ambient air carcinogenic vapors - improved sampling and
analytical technicques and field studies. Report Nos. NTIS PB-297932 and EPA-600/2-79-081.
Prepared for the U.S. Environmental Protection Agency by the Research Triangle Institute,
Research Triangle Park, NC. (As cited in Brodzinsky and Singh, 1983).
Pellizzari, E.D., T.D. Hartwell, B.S.H. Harris, III, R.D. Waddell, D A.
Whitaker and M.D. Erickson. 1982. Purgeable organic compounds in mother's milk. Bull.
Environ. Contam. Toxicol. 28:322-328. (As cited in U.S. EPA, 1994b.)
Pellizzari, E.D., R.A. Zweidinger, and M.D. Erickson. 1978. Environmental monitoring near
industrial sites: Brominated chemicals Part I. Report Nos. NTIS PB286484 and EPA-560/6-78-
002. Prepared for the U.S. Environmental Protection Agency by the Research Triangle Institute,
Research Triangle Park, NC. (As cited in Brodzinsky and Singh, 1983).
Potter, C.L., L.W. Chang, A.B. DeAngelo, and F.B. Daniel. 1996. Effects of four
trihalomethanes on DNA strand breaks, renal hyaline droplet formation, and serum testosterone in
male F-344 rats. Cancer letters. 106(2):235-242.
Roldan-Arjona, T. and C. Pueyo. 1993. Mutagenic and lethal effects of halogenated methanes in
the Ara test of Salmonella typhimurium: Quantitative relationship with chemical reactivity.
Mutagenesis 8:127-131.
Reif, J., M.C. Hatch, M. Bracken, L.B. Holmes, B.A. Schwetz, and P.C. Singer. 1996.
Reproductive and developmental effects of disinfection by-products in drinking water. Environ.
Health Perspect. 104:1056-1061.
Draft -Do not cite or quote
IX- 15
February 20, 2002
-------
Raucy, J.L. 1995. Risk Assessment: toxicity from chemical exposure resulting from expression
ofCYP2E1. Toxicology. 105:217-223.
Ries, L.A.G., M.A. Smith, J.G. Gurney, M. Linet, T. Tamra, J.L. Young, and G.R. Bunin (eds).
1999. Cancer Incidence and Survival Among Children and Adolescents: United States SEER
Program 1975-1995. Bethesda, MD: National Cancer Institute, SEER Program. NIH
Publication No. 99-4649. Available on-line at http://www-seer.ims.nci.nih.gov.
Ronis, M.J.J., K.O. Lindros, and M. Ingelman-Sundberg. 1996. The CYP2E Subfamily. In:
Ioannides, C., ed. Cytochromes P450: Metabolic and toxicological aspects. Boca Raton: CRC
Press, pp. 211-239.
Roth, O. 1904. Zeitschr. f. Medizinabeamte. 17:233. (As cited in U.S. EPA, 1994b.)
Ruddick, J. A., D.C. Villeneuve and I. Chu. 1983. A teratological assessment of four
trihalomethanes in the rat. J. Environ. Sci. Health. 1318(3):333-349.
Savitz, D. A., K. W. Andrews and L. M. Pastore. 1995. Drinking water and pregnancy outcome
in central North Carolina: source, amount, and trihalomethane levels. Environ. Health Perspect.
103(6): 592-6.
Sekihashi, K., A. Yamamoto, Y. Matsumura, S. Ueno, M. Watanabe-Akanuma, F. Kassie, S.
Knasmuller, S. Tsuda, Y.F. Sasaki. 2002. Comparative investigation of multiple organs of mice
and rats in the comet assay. Mutat. Res. 517:53-75.
Shikiya, J., G. Tsou, J. Kowalski and F. Leh. 1984. Ambient monitoring of selected halogenated
hydrocarbons and benzene in the California south coast air basin. Proc-APCA 77th Annual Mtg.,
Vol. 1, pp. 84-1.1. (As cited in U.S. EPA, 1994b.)
Simmon, V.F. and R.G. Tardiff. 1978. The mutagenic activity of halogenated compounds found
in chlorinated drinking water. In: Jolley, R.H., W.A. Brungs and R.B. Cumming, eds. Water
chlorination: Environmental impacts and health effects, Vol. 2. Ann Arbor, MI: Ann Arbor
Science, pp. 417-431. (As cited in U.S. EPA, 1994b.)
Sobti, R.C. 1984. Sister chromatid exchange induction potential of the halogenated
hydrocarbons produced during water chlorination. Chromosome Information Service No. 37
pp 17-19. (As cited in U.S. EPA, 1994b.)
Sofuni, T., M. Honma, M. Hayashi, H. Shimada, N. Tanaka, S. Wakuri, T. Awogi, K.I.
Yamamoto, Y. Nishi, and M. Nakadate. 1996. Detection of in vitro clastagens and spindle
poisons by the mouse lymphoma assay using the micro well method: interim report of an
international collaborative study. Mutagenesis. ll(4):349-355.
Squillace, P.J., M.J. Moran, W.W. Lapham, C.V. Price, R.M. Clawges, and J.S. Zogorski. 1999.
Volatile organic compounds in untreated ambient groundwater of the United States, 1985-1995.
Environ. Sci. Technol. 33:4176-4187.
Draft -Do not cite or quote
IX- 16
February 20, 2002
-------
Stevens, J.L. and M.W. Anders. 1979. Metabolism of haloforms to carbon monoxide: III.
Studies on the mechanism of the reaction. Biochem. Pharmacol. 28:3189-3194. (As cited in
U.S. EPA, 1994b.)
Stevens, J.L. and M.W. Anders. 1981. Metabolism of haloforms to carbon monoxide: IV.
Studies on the reaction mechanism in vivo. Chem. Biol. Interact. 37:365-374. (As cited in U.S.
EPA, 1994b.)
Stocker, K.J., J. Statham, W.R. Howard, et al. 1997. Assessment of the potential in vivo
genotoxicity of three trihalomethanes: chlorodibromomethane, bromodichloromethane and
bromoform. Mutagenesis. 12(3): 169-173.
Struba, R.J. 1979. Cancer and drinking water quality. Ph.D. Thesis. University of North
Carolina, Chapel Hill, North Carolina. 156 pp. Available from: Ann Arbor, MI: University
Microfilms International. Publication No. 8022514. (As cited in U.S. EPA, 1994b.)
Swan, S. H., K. Waller, B. Hopkins, G. Windham, L. Fenster, C. Schaefer and R.R. Neutra.
1998. A prospective study of spontaneous abortion: relation to amount and source of drinking
water consumed in early pregnancy. Epidemiology. 9(2): 126-3 3.
Symons, J.M., T.A. Bellar, J.K. Carswell, J. DeMarco, K.L. Kropp, G.G. Robeck, D.R. Seeger,
C.J. Slocum, B.L. Smith and A.A. Stevens. 1975. National organics reconnaissance survey for
halogenated organics. J. Amer. Water Works Assoc. 67(11):634-647. (As cited in U.S. EPA,
1994b.)
Templin, M.V., J.L. Larson, B.E. Butterworth, K.C. Jamison, J.R. Leininger, S. Mery, K.T.
Morgan, B.A. Wong, and D.C. Wolf. 1996. A 90-day chloroform inhalation study in F-344 rats:
Profile of toxicity and relevance to cancer studies. Fundam. Appl. Toxicol. 32:109-125.
Testai, E., V. De Curtis, S. Gemma, L. Fabrizi, P Gervasi, L. Vittozzi. 1996. The role of
different cytochrome P450 isoforms in in vitro chloroform metabolism. J. Biochem. Toxicol.
11:305-312.
Thier, R., J.B. Taylor, S.E. Pemble, W.G. Humphreys, M. Persmark, B. Ketterer, and F.P.
Guengerich. 1993. Expressions of mammalian glutathione S-transferase 5-5 in Salmonella
typhimurium TA1535 leads to base-pair mutations upon exposure to dihalomethanes. Proc. Natl.
Acad. Sci. USA 90:8576-8580.
Theiss, J.C., G.D. Stoner, M.B. Shimkin and E.K. Weisburger. 1977. Test for carcinogenicity of
organic contaminants of United States drinking waters by pulmonary tumor response in strain A
mice. Cancer Res. 37:2717. (As cited in U.S. EPA, 1994b.)
Thornton-Manning, J.R., P. Gao, P.D. Lilly and R.A. Pegram. 1993. Acute
bromodichloromethane toxicity in rats pretreated with cytochrome P450 inducers and inhibitors.
The Toxicologist. 13:361 (Meeting Abstract).
Draft -Do not cite or quote
IX- 17
February 20, 2002
-------
Thornton-Manning, J.R., J.C. Seely and R.A. Pegram. 1994. Toxicity of bromodichloromethane
in female rats and mice after repeated oral dosing. Toxicology. 94(l-3):3-18.
Tobe, M., Y. Suzuki, K. Aida, H. Yoshimoto, K. Takada, J. Kadoma, K. Kobayashi, Y. Vehida,
K. Anda, Y. Nakaji, Y. Suzuki and M. Sato. 1982. Studies on the chronic oral toxicity of
tribromomethane, dibromochloromethane and bromodichloromethane. Unpublished intraagency
report to the National Institute of Hygienic Sciences. Tokyo, Japan: Tokyo Medical and Dental
University. (As cited in IRIS, 1992, U.S. EPA, 1994b.)
Tomasi, A., E. Albano, F. Biasi, T.F. Slater, V. Vannini and M.V. Diazano. 1985. Activation of
chloroform and related trihalomethanes to free radical intermediates in isolated hepatocytes and in
the rat in vivo as detected by the ESR-Spin-trapping technique. Chem. Biol. Interact. 55:303-
316. (As cited in U.S. EPA, 1994b.)
Torti, V.R., A.J. Cobb, J.I. Everitt, M.W. Marshall, G.A. Boorman, and B.E. Butterworth. 2001.
Nephrotoxicity and hepatotoxicity induced by inhaled bromodichloromethane in wild-type and
p53-heterozygous mice. Toxicol. Sci. 64:269-280.
Toyoda, M., K. Takagi, A. Tsurumizu, Y. Saito, H. Kikawa, T. Kawamura and Y. Sakabe. 1990.
Estimation of daily intake of four kinds of trihalomethane by Japanese housewives. J. Food Saf.
10(3):191-200.
Tumasonis, C.F., D.N. McMartin and B. Bush. 1987. Toxicity of chloroform and
bromodichloromethane when administered over a lifetime in rats. J. Environ. Pathol. Toxicol
Oncol. 7:55-64. (As cited in U.S. EPA, 1994b.)
Uhler, A.D. and G.W. Diachenko. 1987. Volatile halocarbons in process water and processed
foods. Bull. Environ. Cont. Toxic. 39(4):601-607. (As cited in U.S. EPA, 1994b.)
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drinking water. Report to
congress, Washington, D.C.: U.S. EPA. EPA-56014-75-005. PB 260961. (As cited in U.S.
EPA, 1994b.)
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality criteria for
halomethanes. Contract no. EPA-440/5-80-051. Washington, DC: U.S. Environmental
Protection Agency, Office of Water Regulations and Standards. (As cited in U.S. EPA, 1994b.)
U.S. EPA. 1985. U.S. Environmental Protection Agency. Health and environmental effects
profile for bromodichloromethane. U.S. Environmental Protection Agency, Environmental
Criteria and Assessment Office, Office of Health and Environmental Assessment, Cincinnati, OH.
June. p. 25. (As cited in U.S. EPA, 1994b.)
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogenic risk
assessment. Fed. Reg. 51(185):33992-34003.
Draft -Do not cite or quote
IX- 18
February 20, 2002
-------
U.S. EPA. 1987. U.S. Environmental Protection Agency. Health effects assessment for
trihalogenated methanes. EPA/600/8-88/059.
U.S. EPA. 1989a. U.S. Environmental Protection Agency. Final report for disinfection
by-products in United States drinking waters. U.S. Environmental Protection Agency,
Association of Metropolitan Water Agencies (AMWA), and the California Department of Health
Services (CDHS). November, pp. 5-1 to 5-35.
U.S. EPA. 1989b. U.S. Environmental Protection Agency. Final report for disinfection
by-products in United States drinking waters: Executive Summary. U.S. Environmental
Protection Agency, Association of Metropolitan Water Agencies (AMWA), and the California
Department of Health Services (CDHS). November. (As cited in U.S. EPA, 1994b.)
U.S. EPA. 1991. U.S. Environmental Protection Agency. Technical Support Division
unregulated contaminant database. Technical Support Division (TSD), Office of Ground Water
and Drinking Water, U.S. Environmental Protection Agency.
U.S. EPA. 1992a. U.S. Environmental Protection Agency. Technical Support Division
disinfection by-products field study data base. Technical Support Division (TSD), Office of
Ground Water and Drinking Water, U.S. Environmental Protection Agency. Communication
with Patricia Fair, TSD, Cincinnati, OH. February 14. (As cited in U.S. EPA, 1994b.)
U.S. EPA. 1992b. U.S. Environmental Protection Agency. Draft report: A cross-species scaling
factor for carcinogen risk assessment based on equivalence of mg/kg3/4/day. Federal Register
57(109):24152-24173.
U.S. EPA. 1992c. United States Environmental Protection Agency. Dermal Exposure
Assessment: Principles and Applications. Interim Report. Exposure Assessment Group. Office of
Health and Environmental Assessment. EPA/600/8-91/01 IB. Washington, D.C. January.
U.S. EPA. 1994a. U.S. Environmental Protection Agency. Final draft for the drinking water
criteria document on chlorine. (As cited in U.S. EPA, 1994b.)
U.S. EPA. 1994b. U.S. Environmental Protection Agency. Final draft for the drinking water
criteria document on trihalomethanes. Contract no. EPA/68-C2-0139. Washington, D.C.: U.S.
Environmental Protection Agency Health and Ecological Division, Office of Science and
Technology, Office of Water.
U.S. EPA. 1995. The Use of the Benchmark Dose Approach in Health Risk Assessment. U.S.
Environmental Protection Agency, Office of Research and Development. EPA/63 0/R-94/007.
U.S. EPA. 1996. U.S. Environmental Protection Agency. Proposed Guidelines for Carcinogen
Risk Assessment. Office of Research and Development, Washington, D.C. EPA/600/P-92/003C
EPA/600/P-92/003C.
Draft -Do not cite or quote
IX- 19
February 20, 2002
-------
U.S. EPA. 1998a. U.S. Environmental Protection Agency. Health Risks to Fetuses, Infants, and
Children (Final Stage 1 D/DBP Rule). Washington, D.C.: U.S. Environmental Protection Agency
Health and Ecological Division, Office of Science and Technology, Office of Water. October 29,
1998.
U.S. EPA. 1998b. U.S. Environmental Protection Agency. Quantification of toxicological
effects for brominated trihalomethanes. By ICF Inc., under contract to The Cadmus Group, Inc.
Contract No. 68-C7-0002, Subcontract No. 0002-ICF-l.
U.S. EPA. 1999. U.S. Environmental Protection Agency. Guidelines for Carcinogen Risk
Assessment. National Center for Environmental Assessment, Risk Assessment Forum,
Washington, D.C. NCEA-F-0644. July 1999. Available on-line at:
http://www.epa.gov/ncea/raf/car2sab.htm.
U.S. EPA. 2000a. Help Manual for Benchmark Dose Software Version 1.20. U.S.
Environmental Protection Agency, Office of Research and Development. EPA600/R-00/014F.
February 2000.
U.S. EPA. 2000b. U.S. Environmental Protection Agency. Benchmark Dose Technical
Guidance Document. External Peer Review Draft. October, 2000. EPA/630/R-00/001.
U.S. EPA. 2000c. U.S. Environmental Protection Agency. Methodology for Deriving Ambient
Water Quality Criteria for the Protection of Human Health (2000). Office of Science and
Technology, Office of Water. EPA-822B-00-004. October, 2000.
U.S. EPA. 2000d. U.S. Environmental Protection Agency. Toxicological Review of Chloroform
(CAS No. 67-66-3). EPA/635/R-01/001. October, 2001.
U.S. EPA. 2001a. U.S. Environmental Protection Agency. Stage 2 Occurrence and Exposure
Assessment for Disinfectants and Disinfection By-Products (D/DBPs) in Public Drinking Water
Systems.
U.S. EPA. 2001b. U.S. Environmental Protection Agency. National Rural Water Association
User Database. Maintained by the Cadmus Group.
U.S. FDA. 2000. U.S. Food and Drug Administration. Total Diet Study: Summary of Residues
Found, Ordered by Pesticide; Market Baskets 91-3 - 99-1. September 2000.
Varma, M.M., F.R. Ampy, K. Verma, and W.W. Talbot. 1988. In vitro mutagenicity of water
contaminants in complex mixtures. J. Appl. Toxicol. 8(4):243-248. (As cited in U.S. EPA,
1994b.)
Vecchia, B.E. and A.L. Bunge. 2002. Skin Absorption Databases and Predictive Equations.
Chapter 2, In: Transdermal Drug Delivery Systems, 2nd Edition. Eds. J. Hadgraft and R.H. Guy,
Marcel Dekker. In press.
Draft -Do not cite or quote
IX-20
February 20, 2002
-------
Vieira, I., M. Sonnier, and T. Cresteil. 1996. Developmental expression of CYP2E1 in the
human liver. Hypermethylation control of gene expression during the neonatal period. Eur. J.
Biochem. 238:476-483.
Vieira, I., M. Pasanen, H. Raunio, and T. Creisteil. 1998. Expression of CYP2E1 in human lung
and kidney during development and in full-term placenta: A differential methylation of the gene is
involved in the regulation process. Pharmacol. Toxicol. 83:183-187.
Voronin V.M., A.I. Donchenko, and A. A. Korolev. 1987. Experimental study of the
carcinogenicity of dichlorobromethane and dibromochloromethane which are formed during the
chlorination of water. Gig. Sanit. p. 19-21 (English translation, as cited in IRIS, 1993a).
Wallace, L. 1981. Measurements of volatile organic compounds in breathing-zone air, drinking
water, and exhaled breath (preliminary draft). (As cited in Brodzinsky and Singh, 1983).
Wallace, L., R. Zweidinger, M. Erickson, S. Cooper, D. Whitaker and E. Pellizzari. 1982.
Monitoring individual exposure: Measurements of volatile organic compounds in breathing-zone
air, drinking water, and exhaled breath. Environment International 8:269-282. (As cited in U.S.
EPA, 1994b.)
Wallace, L.A., E.D. Pellizzari, T.D. Hartwell, C. Sparacino, B. Whitmore, L. Sheldon, H. Zelon
and R. Perritt. 1987. The TEAM study: Personal exposures to toxic substances in air drinking
water, and breath of 400 residents of New Jersey, North Carolina and North Dakota. Environ.
Research. 43:290-307.
Wallace, L.A., E.D. Pellizzari, T.D. Hartwell, B. Whitmore, L. Sheldon, H. Zelon and R. Perritt.
1988. The California TEAM study: Breath concentrations and personal exposure to 26 volatile
compounds in air and drinking water of 1988 residents of Los Angeles, Antioch and Pittsburgh,
CA. Atmospheric Environment. 22:2141-2163.
Wallace, L. 1992. Human exposure and body burden for chloroform and other trihalomethanes.
U.S. Environmental Protection Agency, Office of Research and Development, Warrenton,
Virginia. July 28.
Wallace, L. A. 1997. Human exposure and body burden for chloroform and other
trihalomethanes. Crit. Rev. Environ. Sci. Technol. 27:113-194.
Waller, K. S.H. Swan, G.C. Windham, and L. Fenster. 2001. Influence of exposure assessment
methods on risk estimates in an epidemiologic study of total trihalomethane exposure and
spontaneous abortion. J. Expo. Anal. Environ. Epidemiol. 11:522-531.
Waller, K., S.H. Swan, G. DeLorenze, et al. 1998. Trihalomethanes in drinking water and
spontaneous abortion. Epidemiology. 9(2): 134-140.
Wang, P.Y, T. Kaneko, H. Tsukada and A. Sato. 1994. Dose and route dependency of
metabolism and toxicity of chloroform in ethanol-treated rats. Arch. Toxicol. 69(1): 18-23.
Draft -Do not cite or quote
IX-21
February 20, 2002
-------
Weisel, C. P., H. Kim, P. Haltmeier, and J.B.Klotz. 1999. Exposure estimates to disinfection by-
products of chlorinated drinking water. Environ. Health Perspect. 107: 103-110.
Westrick, J.J., J.W. Mello and R.F. Thomas. 1983. The ground water supply survey summary of
volatile organic contaminant occurrence data. Cincinnati, OH: U.S. Environmental Protection
Agency, Technical Support Division, Office of Drinking Water and Office of Water. (As cited in
U.S. EPA, 1994b.)
Windham, G.C., K. Waller, M. Anderson, L. Fenster, P. Mendola, and S. Swan. 2003.
Chlorination by-products in drinking water and menstrual cycle function. Env. Health Perspect.
In press.
Windholz, M., ed. 1976. The Merck Index. Rathway, NJ: Merck &Co., Inc. (As cited in U.S.
EPA, 1994b).
Wolf, C.R., D. Mansuy, W. Nastainczyk, G. Deutschmann and V. Ullrich. 1977. The reduction
of polyhalogenated methanes by liver microsomal cytochrome P-450. Mol. Pharmacol. 13:698-
705. (As cited in U.S. EPA, 1994b.)
Yang, C-Y., B-H. Cheng, S-S. Tsai, et al. 2000. Association between chlorination of drinking
water and adverse pregnancy outcome in Taiwan. Environ. Health Perspect. 108:765-768.
Young, T.B., M.S. Kanarek and A.A. Tslatis. 1981. Epidemiologic study of drinking water
chlorination and Wisconsin female cancer mortality. J. Natl. Cancer Inst. 67:1191-1198. (As
cited in U.S. EPA, 1994b.)
Young, T.B., D.A. Wolf and M.S. Kanarek. 1987. Case-control study of colon cancer and
drinking water trihalomethanes in Wisconsin. Int. J. Epidemiol. 16(2): 190-197. (As cited in U.S.
EPA, 1994b.)
Zeiger, E. 1990. Mutagenicity of 42 chemicals in Salmonella. Environ. Mol. Mutagen. Suppl.
16(Supplement 18):32-54. (As cited in U.S. EPA, 1994b.)
Zhao, G., amd J.W. Allis. 2002. Kinetics of bromodichloromethane metabolism by cytochrome
P450 isoenzymes in human liver microsomes. Chem. Biol. Interact. 140:155-168.
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APPENDIX A
BENCHMARK DOSE MODELING OF HEALTH EFFECTS ENDPOINTS FOR THE
BROMINATED TRIHALOMETHANES: BROMODICHLOROMETHANE,
DIBROMOCHLOROMETHANE, AND BROMOFORM
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February 20, 2002
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A. INTRODUCTION
The limitations of the NOAEL/LOAEL approach as the basis for estimating thresholds of
toxic effect are well-documented (e.g., U.S. EPA, 1995, 2001b). These limitations include:
1) the slope of the dose-response plays little role in determining the NOAEL;
2) the NOAEL (or LOAEL) is limited to the doses tested experimentally;
3) the determination of the NOAEL is based on scientific judgement, and is subject to
inconsistency;
4) experiments using fewer animals tend to produce larger NOAELs, and as a result may
produce larger health advisories (HAs) or reference doses (RfDs) (U.S. EPA, 1995,
2001b) that may not be sufficiently protective of human health.
In contrast, benchmark doses (BMDs) are not limited to the experimental doses,
appropriately reflect the sample size, and can be defined in a statistically consistent manner. In
light of these considerations, it is becoming common practice to conduct assessments by
performing BMD modeling for key endpoints, in addition to identification of NOAELs and
LOAELs.
This document describes the analysis of the data relevant to the development of the One-
day, Ten-day, and Longer-term Health Advisories (HAs) for bromodichloromethane (BDCM),
dibromochloromethane (DBCM), and bromoform. Available data of appropriate duration were
analyzed and the implications of the calculated benchmark doses for the development of HAs
were considered. Comparisons of the resulting health advisories with existing values are also
made. Developmental and reproductive toxicity studies were also considered when effects were
seen at doses comparable to or lower than those causing systemic toxicity in subchronic or
chronic studies. The data modeled in support of HA development were also used in derviation of
the reference doses for the three brominated trihalomethanes.
B. SELECTION OF STUDIES AND ENDPOINTS FOR MODELING
The large number of candidate data sets for BMD modeling -required development of a
data prioritization system. The available studies were first reviewed for endpoints and data sets
appropriate for BMD modeling. Priority for modeling was given to those endpoints that showed
the greatest toxicological relevance (e.g., developmental/reproductive endpoints, target organ
histopathology) and ease of interpretation. Ease of interpretation refers to the ability to
characterize the response as adverse and to translate this into an appropriate response level for
input into the BMDS program. In addition, endpoints for which the LOAEL was less than ten
times the lowest LOAEL observed in the category (e.g., 1-day, 10-day, and Longer term HAs,
RfD etc.) were given priority for modeling. Data considered for BMD and general criteria for
selection are listed in Tables A-l, A-2, and A-3.
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Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Candidate Studies for Derivation of the One-day HA
Lilly et al.
(1994)
Rat
M
Gavage
(oil)
6
0
200
400
Single
Dose
Kidney wt
200
400
Yes
Low
No
Model data for
aqueous vehicle
from this study
(see below)
Rel. kidney wt.
200
400
Yes
Low
No
Serum & urine
chem :
Serum AST,
LDH, ALT,
Creatinine, BUN
Urine pH,
osmolality
200
400
Yes
Generally
low
No
Kidney
histopath
minimal renal
tubule
degeneration
and necrosis
200
Yes
High
No
Liver histopath
minimal
vacuolar
degeneration
200
400
Yes
High
No
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A - 3
February 20, 2002
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Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Lilly et al.
(1994)
Rat
M
Gavage
(aqueous)
6
0
200
400
Single
Dose
Body wt.
200
400
Yes
Moderate
No
Model midzonal
vacuolar
degeneration (48
hr) and renal
tubule
degeneration (48
hr)
Table 2, p. 135
Model SDH as test
Liver wt
-
200
Yes
Low
No
Rel. liver wt.
-
200
Yes
Low
No
Kidney weights
200
400
Yes
Low
No
Rel kidney wt
-
400 NS
Yes
Low
No
Serum & urine
chem
Serum AST,
LDH, ALT,
Creatinine,
BUN, urine pH,
osmolality
200
400
Yes
Generally
low
No
Kidney
histopath
-
200
Yes
High
Yos
Liver histopath
200
400
Yes
High
Yos
Draft -Do not cite or quote
A - 4
February 20, 2002
-------
Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Lilly et al.
(1997)
Rat
M
Gavage
(aqueous)
5
0
123
Single
Dose
Body wt 48 hr
post
328
492 (for
>10%)
Yes
Low
No
Model SDH 24 hr
post as test
164
246
328
492
Liver wt 48 hr
post
164
246
Yes
Low
No
Rel liver wt 48
hr post
328
492
Yes
Low
No
Kidney wt 24 hr
post
246
328
Yes
Low
No
Rel kidney 24 hr
post
164
246
Yes
Low
No
Rel kidney 48 hr
post
328
492
Yes
Low
No
Serum / urine
chemist SDH 24
hr post
-
123
Yes
Moderate
No
Keegan et
al.
(1998)
Rat
Gavage
(aqueous)
0
21
31
41
Single
Dose
Body wt
82
123
Yes
Moderate
No
Test model SDH-
24 hrs control
group means in
Table 2 as test
Liver wt.
41
82
Yes
Low
No
82
123
Rel kidney wt
123
164
Yes
Low
No
164
246
Serum chemistry
(elevated ALT,
AST, and SDH
activities)
41
82
Yes
Moderate
No
Draft -Do not cite or quote
A - 5 February 20, 2002
-------
Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
French et
al.
Rat
F
Gavage
(aqueous)
3-6
0 (water)
0
(emulph)
75
150
300
5 days
Body wt.
150
300
(FEL)
Yes
Low
(at FEL)
No
Most effects
occurred at FEL
(1999)
Spleen wt.
150
300
Yes
Low
No
PHA is sig. effect
at dose below
FEL. however,
Thymus wt.
150
300
Yes
Low
No
Rel. thymus wt.
150
300
Yes
Low
No
vehicle alone
caused significant
MLNC prolif
ConA
150
300
Yes
Low
No
increase in this
endpoint
MLNC prolif
PHA
75
150
Yes
?
No
Thornton-
Manning
et al.
(1994)
Rat
F
Gavage
(aqueous)
4-6
0
75
150
300
5 days
Body wt
150
300
Yes
Moderate
No
Model liver histo
path (centrilobul.
vacuolar degener.)
Table 4; p. 11
Model kidney
histopath (renal
tubule vacuolar
Liver wt
75
150
Yes
Low
No
Rel. liver wt.
75
150
Yes
Low
No
Kidney wt
75
150
Yes
Low
No
Rel. kidney wt.
75
150
Yes
Low
No
degeneration and
regeneration)
Serum chemistry
(hepatotoxicity)
75
150
Yes
Moderate
No
Table 5, p. 13
Serum chemistry
(renal toxicity)
150
300
Yes
Low
No
model SDH as test
Draft -Do not cite or quote
A - 6 February 20, 2002
-------
Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Liver/kidney
histopath
(mild to
moderate
centrilobular
hepatocell.
vacuolar
degeneration,
mild renal
tubule vacuolar
degener.)
75
150
Yes
High
Yes
Thornton-
Manning
et al.
(1994)
Mouse
F
Gavage
(aqueous)
5-6
0
75
150
5 days
Liver wt
75
150
Yes
Low
No
Serum chemistry
SDH
-
75
Yes
Moderate
No
Draft -Do not cite or quote
A - 7
February 20, 2002
-------
Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
Comments
Candidate Studies for Derivation of the Ten-day HA
Aida et al.
(1992a)
Rat
M
F
Feed
7
7
0
21
62
189 M
204 F
1 month
Body wt (M)
62
189
Yes
Moderate
No
Model Liver cell
vacuol. in females
Table 8 p. 129
Liver wt (M)
62
189
Yes
Low
No
Kidney wt (M)
62
189
Yes
Low
No
Body wt. (F)
62
204
Yes
Moderate
No
Rel. liver wt. (F)
62
204
Yes
Low
No
Liver histopath
(M)
62
189
Yes
High
No
Liver histopath
(F)
21
62
Yes
High
Yes
Chu et al.
(1982a)
Rat
M
Drinking
water
10
0
0.8
8.0
68
28 days
Clinical signs
serum chemistry
histology
68
-
Yes
Low
No
Lack of effect; No
data selected for
modelling
Condie et
al.
(1983)
Mouse
Gavage
(oil)
9-
10
0
37
74
148
14 days
Serum enzymes
[elevated
SPGT/ALT]
74
148
Yes
Moderate
No
Model kidney
histopath
(Epithelial
hyperplas.);
Table 4, p. 571
Liver histopath
(centrilob. pallor)
Table 5, p. 572
Decreased PAH
uptake in vitro
37
74
Yes
Low
No
Liver Histopath
37?
74
Yes
Mod. - High
Yes
Kidney
Histopath
74
148
Yes
High
Yes
Draft -Do not cite or quote
A - 8 February 20, 2002
-------
Table A-l Candidate Studies and Data for BMD Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Melnick et
al.
(1998)
Mouse
Gavage
(oil)
10
0
75/54
150/107
326/233
3 weeks
(5 d/wk)
Liver wt
75
150
Yes
Low
No
Model hepatocyte
hydropic degener.
Fig. 4, p. 142
Serum chem
-
75
Yes
Moderate
No
Liver histopath
75
150
Yes
High
Yes
Labeling index
75
150
Yes
Moderate
No
Munson et
al.
(1982)
Mouse
Gavage
0
14 days
Body wt (M)
125
250
Yes
Moderate
No
M
F
(aq)
8-
12
50
125
250
Rel liver wt (M)
50
125
Yes
Low
No
Spleen wt (M)
125
250
Yes
Low
No
Serum chem (M)
SPGT
SGOT
125
250
Yes
Low -
moderate
No
Hematology
(M)
125
250
Yes
Low
No
Hemagglut (M)
50
125
Yes
Low
No
Body wt (F)
125
250
Yes
Moderate
No
Rel liver wt (F)
50
125
Yes
Low
No
Spleen wt (F)
50
125
Yes
Low
No
Rel Spleen wt
(F)
50
125
Yes
Low
No
Serum chem (F)
SPGT
SGOT
125
250
yes
Low
No
Draft -Do not cite or quote
A - 9 February 20, 2002
-------
Table A-l Candidate Studies and Data for BMP Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Hematology (F)
125
250
Yes
Low
No
AFC/spleen (F)
50
125
Yes
Low-
Moderate?
No
Hemagglutin (F)
125
250
Yes
Low
No
NTP
(1987)
Rat
M
F
Gavage
(oil)
5M
4-
5F
0
38
75
150
300
600
14 days
Body wt (M)
150
300
Yes
Moderate
No
LOAEL for effect
higher than other
for other
endpoints
NTP
(1987)
Mouse
M
F
Gavage
(oil)
5M
4-
5F
0
19
38
75
150
300
14 days
Mortality,
lethargy, gross
renal pathology
75
150
(FEL)
Yes
Low
No
No suitably
sensitive endpoint
NTP
(1998)
Rat
M
F
Drinking
water
6
0
9
38
67
(Grp A
males)
30 days
Liver histopath
9
38
Yes
High
Yes
Hepatocyte indiv.
cell necrosis
Table 2, p.36
Narotsky
et al.
(1997)*
Rat
F
Gavage
(oil)
(aq.)
12-
14
0
25
50
75
Gestation
days 6-15
Full-litter
resorption
25
50
Yes
High
Yes
Model full litter
resorption
(aqueous vehicle)
Fig. 2, incidence
reported in text
above table
Draft -Do not cite or quote
A - 10 February 20, 2002
-------
Table A-l Candidate Studies and Data for BMP Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Bielmeier
et al.
(2001)
Rat
F
Gavage
(aq.)
8-
11
0
75
100
Gestation
day 9
Full-litter
resorption
75
Yes
High
Yes
Model full litter
resorption
Table p. 23 of
manuscript
("Hormone profile
if)
Coffin et
al. (2000)
Mouse
Gavage
(corn oil)
10
0
150
300
11 days
Liver
histopathology
-
150
Yes
High
No
Aida et al. (1992a)
used an additional,
lower dose which
Increased
labeling index
-
200
Yes
Moderate
No
provides more
information about
shape of curve in
low dose region
Increased
relative liver wt.
-
200
Yes
Low
No
for histopath.
effects. No
incidence data for
histopathology.
Candidate Studies for Derivation of the Longer-term HA
NTP
(1987)
Rat
M
F
Gavage
(oil)
9-
10
0/0
19/14
38/27
75/54
150/107
300/214
13 weeks
(5 d/wk)
Body weight
75/54
(M)
150/107
(F)
150/107
(M)
300/
214(F)
Yes
Moderate
No
Data in text on p.
35-36
Histopath effects
occurred only at
FEL
Hepatic and
renal histopath
(M)
150/107
300/
214
Yes
High
No
Draft -Do not cite or quote
A- 11
February 20, 2002
-------
Table A-l Candidate Studies and Data for BMP Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
NTP
(1987)
Mouse
M
Gavage
(oil)
10
0
6.25/4.5
12.5/9
25/18
50/36
100/71
13 weeks
(5 d/wk)
Renal histopath
50/36
100/71
Yes
High
Yes
BMD modelling
conducted by ICF
on data for
focal necrosis of
renal tubular
epithelium in
males
NTP
(1987)
Mouse
F
Gavage
(oil)
10
0
25/18
50/36
100/71
200/142
400/284
13 weeks
(5 d/wk)
Liver histopath
Vacuolated
cytoplasm
50/36
100/71
Yes
High
Yes
Data in text on p.
49
Model vacuolated
cytoplasm
Reproductive and Developmental Studies
Ruddick
et al.
(1983)
Rat
Gavage
(oil)
9-
14
0
50
100
200
GD 6-15
Sternebral
aberrations
100
200
Yes
High
Yes
Model sternebra
variations
Narotsky
et al.
(1997)
Rat
Gavage
(oil)
12-
14
0
25
50
75
GD 6-15
Developmental
Full litter
resorption
25
50
Yes
Moderate
(vehicle)
No
Model aq. data,
from same study
Narotsky
et al.
(1997)
Rat
Gavage
(aq)
12-
14
0
25
50
75
GD 6-15
Developmental
Full litter
resorption
25
50
Yes
High
Yes
This study listed
in table for Longer
term HA.
Maternal
Reduced body
weight gain
-
25
Yes
Moderate to
high
Yes
ccc
(2000a)
Rabbit
Drinking
water
5
0
4.9
13.9
32.3
76.3
Gestation
days 6-29
Reproductive
developmental
endpoints
76
No
Potentially
High
No
No adverse
effects; small
sample size
Draft -Do not cite or quote
A - 12 February 20, 2002
-------
Table A-l Candidate Studies and Data for BMP Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
ccc
(2000b)
Rabbit
Drinking
water
25
0
1.4
13.4
Gestation
days 6-29
Reproductive/
developmental
55
-
-
Potentially
High
No
No adverse repro.
or develop,
effects. Model
35.6
55.3
Maternal
Reduced body
weight gain
13.4
36
Yes
Moderate to
high
Yes
corrected
maternal wt. gain
gd 6-29 as
maternal effect.
CCC
(2000c)
Rat
Drinking
water
10
Females
0 ppm
50 ppm
150 ppm
450 ppm
1350
ppm
Gestation
days 0-21
Reproductive/
developmental
50 ppm
150
Potentially
High
No
Decreased pup wt.
and wt. gain at
doses that caused
parental toxicity;
reliable mg/kg-
day dose could
not be estimated.
CCC
(2000d)
Rat
Drinking
water
25
0.0
2.2
18.4
45.0
82.0
Gestation
days 6-21
Developmental
Reduced
number of
ossification sites
in phalanges and
metatarsals
45
82
Yes
Moderate
No
Reversible
variation
occurring at doses
that cause
maternal toxicity
Maternal
Reduced body
weight gain
18.4
45
Yes
Moderate to
high
Yes
Model body
weight gain for
gestation days 6-7
and 6-9.
Draft -Do not cite or quote
A - 13
February 20, 2002
-------
Table A-l Candidate Studies and Data for BMP Modeling - Bromodichloromethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
(mg/kg-
day)
LOAEL
(mg/kg-
day)
L<10*LLf
Ease of
Interp/Toxi
cologic
Relevance
BMD
model
7
Comments
Bielmeier
et al.
(2001)
Rat
Gavage
(aq.)
8-
11
0
75
100
Gestation
day 9
Full-litter
resorption
75
Yes
High
Yes
Model full litter
resorption
Table p. 23 of
manuscript
("Hormone profile
n")
Study also listed
under Longer-
term HA
¦f L
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dibromochloromethane
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
Candidate Studies for Derivation of the One-day HA - none
Candidate Studies for Derivation of the Ten-day HA
Aida et al.
(1992a)
Rat
M
F
Feed
7
Males
0
18
56
173
Females
0
34
101
332
1 month
Liver wt (M)
56
173
Yes
Moderate
No
Model:
liver
histopath
(liver cell
vacuoliza-
tion) in M
and F
Table 8,
p.129
Rel liver wt (M)
56
173
Yes
Moderate
No
Liver histopath
(M)
56
173
Yes
High
Yes
Body wt (F)
101
332
Yes
Moderate
No
Liver wt (F)
34
101
Yes
Low
No
Rel liver wt (F)
-
34
Yes
Low
No
Rel kidney wt
(F)
101
332
Yes
Low
No
Liver histopath
(F)
101
332
Yes
High
Yes
Chu et al.
(1982a)
Rat
M
Drinking
water
10
0
.7
8.5
68
28 days
-
68
-
-
-
-
No adverse
effects
observed
Draft -Do not cite or quote
A - 15
February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
Condie et
al.
(1983)
Mouse
M
Gavage
(oil)
8-16
0
37
74
147
14 days
Serum SPOT
74
147
Yes
Moderate
No
Model renal
mesangial
hypertrophy
and hepatic
cytoplasmic
vacuolation
Liver histopath
74
147
Yes
High
Yes
Renal histopath
74
147
Yes
High
Yes
Table 5,
p.572
Melnick et
al.
(1998)
Mouse
F
Gavage
(oil)
10
0
50/37
100/71
192/137
417/298
3 weeks
(5 d/wk)
Relative Liver
wt
-
50
Yes
Low
No
Model
Incidence of
hepatocyte
hydropic
degeneration
Table 4,
Serum ALT
100
192
Yes
Moderate
No
Serum SDH
-
50
Yes
Moderate
No
Liver histopath
100
192
Yes
High
Yes
p. 142
Inc. labeling
index
192
417
Yes
High
No
Munson et
al.
(1982)
Mouse
M
F
Gavage
(aq)
8-12
0
50
125
250
14 days
Body wt. (M)
125
150
Yes
Moderate
No
Rel liver wt (M)
50
125
Yes
Low
No
Spleen wt (M)
125
250
Yes
Moderate?
No
Rel spleen wt
(M)
125
250
Yes
Moderate?
No
Hematology -
Fibr (M)
125
250
Yes
Low
No
Serum chem
SGPT (M)
125
250
Yes
Low
No
AFC/Spleen (M)
125
250
Yes
Moderate
No
Draft -Do not cite or quote
A - 16 February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
*AFC/106(M)
50
125
Yes
Moderate
No
Liver wt (F)
125
250
Yes
Low
No
Rel liver wt (F)
50
125
Yes
Low
No
Hematology -
Fibr (F)
125
250
Yes
Low
No
Serum SGPT (F)
125
250
Yes
?
No
AFC/spleen (F)
125
250
Yes
Moderate
No
AFC/106 (F)
50
125
Yes
Moderate
No
NTP
(1985)
Rat
M
F
Gavage
(oil)
5
0
60
125
250
500
1000
14 days
Body wt (M)
250
500 (FEL)
No
Moderate
No
Tables 3 and
4 p.33
Effects
observed
only at
levels where
reduced
survival
occurred:
survival(2/5
and 0/5 for
M and F,
respectively
at the 250
mg/kg-day
LOAEL
Dark'd kid
medulla (M)
250
500
No
Low
No
Mottled liver
(M)
500
1000
No
Low
No
Dark'd kid
medulla (F)
250
500
No
Low
No
Mottled liver (F)
250
500
No
Low
No
Draft -Do not cite or quote
A - 17 February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
NTP
(1985)
Mouse
M
F
Gavage
(oil)
5
0
30
60
14 days
Stomach
nodules (F)
125
250
Yes
Moderate
No
Model
stomach
nodules in M
125
250
500
Stomach
nodules (M)
60
125
Yes
Moderate
Yos
and F
Table 13
p.44
Red'd kid
medulla (F)
250
500
No
Low
No
Renal and
hepatic
effects
Red'd kid
medulla (M)
250
500
No
Low
No
observed
only at
levels where
reduced
Mottled liver
(M)
125
250
Yes
Low
No
survival
observed:
Survival at
500 mg/kg-
day 1/5 and
2/5 for M
and F
respectively
Mottled liver (F)
125
250
Yes
Low
No
Coffin et
al. (2000)
Mouse
F
Gavage
(oil)
10
0
100
300
11 days
Liver
histopathology
-
100
Yes
High
No
Other studies
showing
histopath.
Increased
labeling index
-
100
Yes
Moderate
No
effects used
lower range
of doses. No
incidence
data for
histopathol-
ogy-
Increased
relative liver wt.
-
100
Yes
Low
No
Draft -Do not cite or quote
A-18 February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
NTP
(1996)
Rat
Drinking
water
10
Males
0
4.2
12.4
28.2
Group
A
Females
0
6.3
17.4
46.0
29 days
No clearly
treatment-
related adverse
effects observed
28
Decreased
wt gain
observed in
some
groups, but
effect did
not reach
statistical
significance
Group
B
Females
0
7.1
20
47.8
Draft -Do not cite or quote
A - 19
February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
Candidate Studies for Derivation of the Longer-term HA
Chu et al.
(1982b)
Rat
M
F
Drinking
water
20
Males
0
0.57
6.1
49
224
90 days
Liver histopath -
prevalence (M)
?
?
Yes
High
Yos
Model
incidence
data
Tables 5 and
6
Females
0
0.64
6.9
55
236
Liver histopath -
prevalence (M)
49
224
Yes
High
Yos
"treatment"
results
Daniel et
al.
(1990)
Rat
M
F
Gavage
(oil)
10
0
50
100
200
90 days
Hepatic and
renal lesions
Modeled
previously by
ICF
50
Yes
High
Crump
BMDL10 =
4.2
(kidney
cortex
degeneration
in females)
NTP
(1985)
Rat
Gavage
(oil)
10
0
15
30
60
125
250
13 weeks
(5 d/wk)
hepatic lesions
Modeled
previously by
ICF
30
60
Yes
High
Crump
BMDL10 =
0.93
(liver fatty
metamorpho
sis in males)
Draft -Do not cite or quote
A - 20 February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
NTP
(1985)
Mouse
M
F
Gavage
(oil)
10
0
15
30
60
125
250
13 weeks
(5 d/wk)
Liver histopath
(M)
Kidney
histopath (F)
125
250
Yes
High
No
Occurred
only at
highest dose;
data
provided
only for 0,
125, and 250
mg/kg
doses;
Incidence at
125 0/10 for
all
endpoints;
5/10 for
hepatic vac.
change and
nephropathy
in males;
incidence at
15, 30, and
60 not
examined.
Draft -Do not cite or quote
A - 21
February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
?
Comments
Reproductive and Developmental Studies
Borzelleca
and
Carchman
(1982)
Mouse
M
F
Drinking
Water
10 M
30 F
0
17
171
685
25-27 weeks
Postnatal body
wt.
(cannot be
modeled due to
insufficient data
on number of
litters evaluated)
17
(marginal
)
Yes
High
No
Marginal
LOAEL for
parental
toxicity is 17
mg/kg-day
Ruddick
et al.
(1983)
Rat
F
Gavage
(Corn oil)
9-14
0
50
100
200
g.d. 6-15
"
None
identified
None
identified
"¦
""
No
No clearly
adverse
effect
NTP
(1996)
Rat
M
Drinking
Water
10
4.2
12.4
28.2
29 days
28.2
No
No clearly
adverse
effect on any
reproductive
endpoint at
tested doses
NTP
(1996)
Rat
F
Drinking
Water
10
6.3
17.4
46.0
35 days
46.0
No
No clearly
adverse
effect on any
reproductive
or
development
al endpoint
at tested
doses
NTP
(1996)
Rat
F
Drinking
Water
7.1
20.0
47.8
13
6 days
47.8
No clearly
adverse
effect on any
reprod or
develop
endpoint at
tested doses
Draft -Do not cite or quote
A - 22 February 20, 2002
-------
Table A-2 Candidate Studies and Data for BMD Modeling -Dichlorobromomethane (cont.)
¦f L
-------
Table A-3 Candidate Studies and Data for BMD Modeling - Bromoform
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
7
Comments
Candidate Studies for Derivation of the One-day HA for Bromoform - No suitable studies
Candidate Studies for Derivation of the Ten-day HA for Bromoform
Aida et al.
(1992a)
Rat
M
F
Feed
7
Males
0
62
187
618
Females
0
56
208
728
1 month
Liver histopath
(M)
62
187
Yes
High
Yes
Model liver cell
vacuolization in M
and F
Table 7, p. 128
Serum LDH
56
208
Yes
Low
No
BUN (F)
56
208
Yes
Low
No
Liver histopath.
(F)
56
208
Yes
High
Yes
Chu et al.
(1982a)
Rat
M
Drinking
water
20
0
0.7
8.5
80
28 days
None
80
-
Yes
--
No
No adverse effects
Condie et
al.
(1983)
Mouse
M
Gavage
(oil)
5-
16
0
72
145
289
14 days
Renal slice
uptake PAH
145
289
Yes
Low
No
Model Liver
histopath:
centrilobular pallor
Model kidney
histopath:
mesangial
nephrosis
Table 4, p.571
Renal Histopath
145
289
Yes
High
Yes
Liver histopath
145
289
Yes
High
Yes
SGPT activity
145
289
Yes
Low
No
Draft -Do not cite or quote
A - 24 February 20, 2002
-------
Table A-3 Candidate Studies and Data for BMD Modeling - Bromoform (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
7
Comments
Melnick et
al.
Mouse
F
Gavage
(oil)
10
0
200
3 weeks
(5 d/wk)
Rel Liver wt
200
500
Yes
Low
No
Stat sign increase
in rel liver wt at
(1998)
500
Serum chemistry
ALT
200
500
Yes
?
No
200 - reported to
be about 17% in
text. Value of 200
Serum
Chemistry SDH
200
500
Yes
?
?
for NOAEL based
on consistency of
effects at higher
dose per Mantus
Model: liver
hydropic
Liver histopath
200
500
Yes
High
Yes
Labeling index
200
500
Yes
Moderate
No
degeneration
Graph p. 140
Coffin et
al. (2000)
Mouse
F
Gavage
(oil)
10
0
200
11 days
Liver histopath.
-
200
Yes
High
No
Other studies with
lower range of
500
Labeling index
-
200
Yes
Moderate
No
dose. No
incidence data for
histopathology.
Draft -Do not cite or quote
A-25
February 20, 2002
-------
Table A-3 Candidate Studies and Data for BMD Modeling - Bromoform (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
7
Comments
Munson et
al.
(1982)
Mouse
M
F
Gavage
(aq.)
7-
12
0
50
125
250
14 days
Liver wt (M)
50
125
Yes
Low
No
Rel liver wt (M)
50
125
Yes
Low
No
Hematology -
Fibr (M)
125
250
Yes
Low
No
*Serum SGOT
(M)
125
250
Yes
Low
No
NTP
(1989a)
Mouse
M
F
Gavage
(oil)
5
Male
0
50
100
200
400
600
Female
0
100
200
400
600
800
14 days
Stomach
nodules (M)
200
400
Yes
Moderate
Yes
Model incidence of
stomach nodules in
males
p. 45
Males:
400 4/5
600 3/5
Females:
600 2/5
800 1/5
Stomach
nodules (F)
400
600
Yes
Moderate
No
NTP
Rat
Gavage
5
0
14 days
Body wt (M)
200
400
Yes
Moderate
No
Possibly model
(1989a)
M
(oil)
100
body weight
F
200
400
p. 36
600
800
Draft -Do not cite or quote
A - 26 February 20, 2002
-------
Table A-3 Candidate Studies and Data for BMD Modeling - Bromoform (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
7
Comments
Candidate Studies for Derivation of the Longer-term HA
Chu et
al.
(1982b)
Rat
M
F
Drinkin
g water
9-
10
Males
0
0.65
6.1
57
218
Female
s
90 days
Liver
Histopath (M)
57
218
Yes
High
Yes
Incidence and
mean severity
score provided
for combined
hepatic lesions.
Model
'treatment'
prevalence for
liver lesions in
M and F
Serum chem
data (LDH)
presented only
for high dose
(Insuff data for
modeling)
0
0.64
6.9
55
283
Liver
Histopath (F)
55
283
Yes
High
Yes
NTP
(1989a)
Rat
M
F
Gavage
(corn
oil)
10
0
12
25
50
100
200
13 weeks
(5 d/wk)
Liver
Histopath.
25
50
(hepati
c
vacuoli
zation)
Yes
High
Previously
calculated
Crump BMDL10
2.65
(hepatic
vacuolization in
male rats)
Draft -Do not cite or quote
A - 27 February 20, 2002
-------
Table A-3 Candidate Studies and Data for BMD Modeling - Bromoform (cont.)
Reference
Species
Sex
Route
n
Doses
Exposure
Duration
Candidate
Endpoints
NOAEL
LOAEL
L<10*LLt
Ease of
Interpreta-
tion/
Toxicologic
Relevance
BMD
model
7
Comments
NTP
(1989a)
Mouse
Gavage
(corn oil)
10
0
25
50
100
200
400
13 weeks
(5 d/wk)
Liver Histopath
100
200
Yes
High
Yos
Model hepatic
vacuolization
Reproductive and Developmental Studies
Ruddick
et al.
(1983)
F
Gavage
(Corn oil)
14-
15
0
50
100
200
gd 6-15
Sternebra
aberrations
50
100
Yes
High
Yos
Dose-dependent
increase in
sternebra
aberrations;
intraparietal
deviations at mid-
and high doses.
Intraparietal
variations
-
-
-
High
No
NTP
(1989b)
Mouse
M
F
Gavage
(oil)
20
20
0
50
100
200
(NOAE
L)
105 days
No adverse
effects at doses
tested
200
High
No
No detectable
effect on fertility,
litters/pair, live
pups/litter;
proportion of live
births, sex of live
pups, or pup body
weight.
¦f L
-------
C. METHODS
Benchmark Dose
The brominated trihalomethane data sets considered for dose-response modeling include
both quantal and continuous endpoints. EPA's Benchmark Dose Software (BMDS) (U.S. EPA,
2000a) was used to accomplish all of the model fitting and estimation of the BMD and lower 95%
confidence limit (BMDL). The methods and models applied to both quantal and continuous
endpoints are presented here.
Quantal Models
Seven of the nine quantal models implemented in the BMDS package were used to
represent the dose-response behavior of the quantal endpoints. Specifically, the models used were
the gamma model, the logistic and log-logistic models, the probit and log-probit models, the
multistage model, and the Weibull model. Two other models, the linear and the quadratic models,
were not fit to the data because they are special cases of both the multistage and the Weibull
models. If the fitting of the multistage or Weibull models resulted in a linear or a quadratic form,
then those result were used; otherwise, the linear or quadratic models would not provide a fit as
good as the multistage or Weibull model and so were not separately obtained.
The equations defining each of these models are presented here (U.S. EPA, 2000a). In all
of the following, P(d) represents the probability of response (i.e., adverse effect) following
exposure to "dose" d. In all of these models, a, (3, and y are model parameters estimated using
maximum likelihood techniques, as described below.
Table A-4 Model Equations used in BMD Calculations for Health Advisories
Model
Equation
Conditions
gamma
P(d) = y + (1 - y) (l/r(a))- J ta-le"'dt
0 < y < 1, P > 0, and a > 1. T(x) is the
gamma function, and the integral runs
from 0 to (3d.
logistic
P(d) = [1 + exp{-(a + (3d)}]"1
(3 > 0
log-
logistic
P(d) = y + (1 - y)[l + exp{-(a +
pln(d))}]-1
The log-logistic model has much the
same form as the logistic model except
when d = 0, in which case P(d) = g. In
this case b > 0, and for the background
parameter y, 0 < y < 1 •
probit
P(d) = (x) is the standard normal cumulative
distribution function and (3 > 0.
Draft -Do not cite or quote
A-29
February 20, 2002
-------
log-probit
P(d) = y + (1 - y) • 0(a + (31n(d))
The log-probit model has a form similar
to the probit model except when d = 0,
in which case P(d) = y. Here 0 < y < 1,
and (3 > 1
multistage
model
P(d) = Y+ (1 - Y) (l - exp{-(M + |32d2
+... + (3„dn)})
all the (3 parameters are restricted to be
nonnegative and 0 < y < 1 • When
applied to the brominated
trihalomethane data sets in these
analyses, the degree of the multistage
model (the highest power on dose in the
above equation, n) was set equal to one
less than the number of dose groups in
the experiment being analyzed.
Weibull
model1
P(d) = Y + (1 - Y)(l - exp{-(3da})
The background parameter y is
restricted to fall between 0 (inclusive)
and 1, and (3 is greater than or equal to
0. For these analyses, the parameter a is
constrained to be greater than or equal
to l.1
'The linear model is a special case of the Weibull model obtained by fixing the parameter a equal to 1. The
quadratic model is a special case of the Weibull model obtained by fixing the parameter a equal to 2.
When fitting all of the above-mentioned quantal models, maximum likelihood methods
were used to estimate the parameters of the models. That method maximizes the log-transformed
likelihood of obtaining the observed data, which is (except for an additive constant) given by
L = E K • ln{P(d,)> + (N, - nj ln{ 1 - P(d,)>]
where the sum runs over i from 1 to k (the number of dose groups), and for group i, d; is the dose
(exposure level), N; is the number of individuals tested, and n; is the number of individuals
responding (U.S. EPA, 2000a).
Continuous Models
The continuous endpoints of interest with respect to brominated trihalomethanes toxicity
were quantitatively summarized by group means and measures of variability (standard errors or
standard deviations). The models used to represent the dose-response behavior of those
continuous endpoints are those implemented in EPA's Benchmark Dose Software (U.S. EPA,
Draft -Do not cite or quote
A-30
February 20, 2002
-------
2000a). These models were the power model, the Hill model, and the polynomial model. These
mathematical models fit to the data are defined here. In all cases, |i(d) indicates the mean of the
response variable following exposure to "dose" d.
The power model is represented by the equation
|i(d) = y + (3da
where the parameter a is restricted to be nonnegative. [The linear model is obtained when a is
fixed at a value of 1. The linear model was not separately fit to the data; if the result of fitting the
power model does not result in the linear form, a = 1, then the linear model does not fit as well as
the more general power model, by definition.]
The Hill model is given by the following equation:
|i(d) = y+(vdn)/(dn + kn))
where the parameters n and k are restricted to be positive. Because the Hill model has four
parameters to be estimated (y, v, n, and k), the power n was fixed equal to 1 when the model was
fit to data sets with only three dose groups, so that the number of estimated parameters did not
exceed the number of data points.
The polynomial model is defined as
^(d) = (30 + (31d + ... + (3„dn
where the degree of the polynomial, n, was set equal to one less than the number of dose groups
in the experiment being analyzed. Note that U.S. EPA (2000a) recommends the use of the most
parsimonious model that provides an adequate fit to the data. It may appear that use of a
polynomial model with degree equal to one less than the number of dose groups would not yield
the most parsimonious model. However, allowing the model to have that degree is not the same
as forcing the model to have that degree; in the model fitting, if fewer parameters (e.g., a lower
degree polynomial) is adequate and consistent with the data, then the fitting will reflect that fact
and a more parsimonious model will be the result. For these analyses, the values of the (3
parameters allowed to be estimated were constrained to be either all nonnegative or all
nonpositive (as dictated by the data set being modeled, i.e., nonnegative if the mean response
increased with increasing dose or nonpositive if the mean response decreased with increasing
dose).
In the case of continuous endpoints, one must assume something about the distribution of
individual observations around the dose-specific mean values defined by the above models. The
assumptions imposed by BMDS were used in this analysis: individual observations were assumed
to vary normally around the means with variances given by the following equation:
o;2 = o2 [^(di)]p
Draft -Do not cite or quote
A - 31
February 20, 2002
-------
where both o2 and p were parameters estimated by the model.
Given those assumptions about variation around the means, maximum likelihood methods
were applied to estimate all of the parameters, where the log-likelihood to be maximized is
(except for an additive constant) given by
L = L [(Ni/2) ln(oi2) + (N, - 1 )s2/2g2 + N1{m1 - ^d,)}2^,2]
where N; is the number of individuals in group i exposed to dose d;, and m; and s; are the observed
mean and standard deviation for that group. The summation runs over i from 1 to k (the number
of dose groups).
Goodness of Fit Analyses
For the quantal models, goodness of fit was determined by the modeling software using
the chi-square test. This test is based on sums of squared differences between observed and
predicted numbers of responders. The degrees of freedom for the chi-square test statistic are
equal to the number of dose groups minus the number of parameters fit by the method of
maximum likelihood (ignoring those parameters that are estimated to be equal to one of the
bounds defining their constraints — see the discussion above about constraints imposed on the
model parameters). When the number of parameters estimated equals the number of dose groups,
there are no degrees of freedom for a statistical evaluation of fit.
For the continuous models, goodness of fit was determined based on a likelihood ratio
statistic. In particular, the maximized log-likelihood associated with the fitted model was
compared to the log-likelihood maximized with each dose group considered to have a mean and
variance completely independent of the means and variances of the other dose groups. It is
always the case that the latter log-likelihood will be at least as great as the model-associated log-
likelihood, but if the model does a "reasonable" job of fitting the data, the difference between the
two log-likelihoods will not be too great. A formal statistical test reflecting this idea uses the fact
that twice the difference in the log-likelihoods is distributed as a chi-square random variable. The
degrees of freedom associated with that chi-squared test statistic are equal to the difference
between the number of parameters fit by the model (including the parameters o2 and p defining
how variances change as a function of mean response level) and twice the number of dose groups
(which is equal to the number of parameters estimated by the "model" assuming independence of
dose group means and variances).
Visual fit, particularly in the low-dose region, was assessed for models that had acceptable
global goodness-of-fit. Acceptable global goodness of fit was either a p-value greater than or
equal to 0.1, or a perfect fit when there were no degrees of freedom for a statistical test of fit.
Choice of 0.1 is consistent with current U.S. EPA guidance for BMD modelling (U.S. EPA,
2000b). Local fit was evaluated visually on the graphic output, by comparing the observed and
estimated results at each data point.
Draft -Do not cite or quote
A-32
February 20, 2002
-------
Goodness-of-fit statistics are not designed to compare different models, particularly if the
different models have different numbers of parameters. Within a family of models, adding
parameters generally improves the fit. BMDS reports the Akaike Information Criterion (AIC) to
aid in comparing the fit of different models. The AIC is defined as -2L+2p, where L is the log-
likelihood at the maximum likelihood estimates for the parameters, and p is the number of model
parameters estimated. When comparing the fit of two or more models to a single data set, the
model with the lesser AIC was considered to provide a superior fit.
Definition of the BMR and Corresponding BMP and BMDL
For all of the quantal endpoints analyzed here, the BMDs and BMDLs were defined based
on BMRs of 5% and 10% extra risk. BMDLs were defined as the 95% lower bound on the
corresponding BMD estimates. Confidence bounds were calculated by BMDS using a likelihood
profile method.
Although the 10% response level was selected as the "point of departure" for all the
quantal endpoints analyzed here, we have chosen to follow standard practice and include results
for both the 5% and 10% level of response. In some cases (see discussions below), a comparison
of the 5% and 10% results gives clues about problems with some of the models. In general, we
have included both for completeness, as there is no current consensus concerning the most
appropriate point of departure except in some particular cases (e.g., use of 5% risk for
developmental toxicity tests where the nesting of effects has been modeled using models
specifically designed for such experimental designs).
For the continuous models, BMDs were implicitly defined as follows:
||i(BMD)-|i(0)| = 8 Oj
where o, is the model-estimated standard deviation in the control group. In other words, the
BMR was defined as a change in mean corresponding to some multiplicative factor of the control
group standard deviation.
The value of 8 used in this analysis was 1.1. This value was chosen based on the work of
Crump (1995), who showed that this choice corresponded to an additional risk of 10% when the
background response rate was assumed to be 1%, with normal variation around the mean (and
constant standard deviation). Although the current analyses allowed for nonconstant standard
deviations and estimated extra risk, while the Crump (1995) comparison was based on constant
standard deviations and additional risk, the values of 1.1 was used for two reasons. First, the
difference between additional and extra risk is small when the background rate is 1% or less, so
that the change from additional to extra risk will have minimal impact on the correspondences
proven by Crump (1995). Second, there can be no such generic, a priori correspondences when
standard deviations are allowed to vary in a manner determined only after the model fitting is
accomplished. Thus, to avoid data set- and model-specific choices for 8, the correspondences
proven by Crump (1995) can be used as the best available, consistent definition of the benchmark
Draft -Do not cite or quote
A-33
February 20, 2002
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response. The definition of the BMR as a change in mean of 1.1 times the control standard
deviation is very close to the default value of 1 standard deviation recommended by recent draft
EPA guidelines (U.S. EPA, 2000b). In the following, BMDs and BMDLs corresponding to 8 =
1.1 are denoted BMD10 and BMDL10, because of the just-noted association of that value of 8 with
10% risk.
As for the quantal models, for all of the continuous models BMDLs were defined as the
95% lower bound on the corresponding BMD. Confidence intervals were calculated using a
profile likelihood method.
Choice of BMDL
The following guidance was followed with regard to the choice of the BMDL to use as a
point of departure for calculation of a health advisory. This guidance is consistent with
recommendations in U.S. EPA (2000b). For each endpoint, the following procedure is
recommended:
1. Models with an unacceptable fit (including consideration of local fit in the low-dose
region) are excluded. Visual fit, particularly in the low-dose region, was assessed for
models that had acceptable global goodness-of-fit.
2. If the BMDL values for the remaining models for a given endpoint are within a factor
of 3, no model dependence is assumed, and the models are considered
indistinguishable in the context of the precision of the methods. The models are then
ranked according to the AIC, and the model with the lowest AIC is chosen as the basis
for the BMDL.
3. If the BMDL values are not within a factor of 3, some model dependence is assumed,
and the lowest BMDL is selected as a reasonable conservative estimate, unless it is an
outlier compared to the results from all of the other models. Note that when outliers
are removed, the remaining BMDLs may then be within a factor of 3, and so the
criteria given in item 2. would be applied.
4. The BMDL values from all modeled endpoints are compared, along with any NOAELs
or LOAELs from data sets that were not amenable to modeling, and the lowest
NOAEL or BMDL is chosen.
5. Models with an unacceptable fit (including consideration of local fit in the low-dose
region) are excluded.
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D. MODELING RESULTS
1. Bromodichloromethane
The majority of endpoints modeled consisted of dichotomous data. BMDS modeling
results for bromodichloromethane dichotomous endpoints are summarized in Table A-5 below.
Four sets of continuous data were modeled. These results are summarized in Section f below.
Detailed output for each model run is compiled in Appendix B, provided in electronic format on
compact disk.
a. Developmental and Reproductive studies
Three data sets for developmental or reproductive toxicity were modeled. When the data
for full litter resorption (FLR) in rats reported by Bielmeier et al. (2000) were analyzed, the
BMDL results for the log-logistic model were low relative to the corresponding BMD estimates
(compared to the estimates obtained from the other models); the results from the log-logistic
model might be that is considered qualitative outliers. The remaining values are still not within
factor of 3, indicating some model dependence of the results. In any case, the multistage model
was chosen as it gave the smallest value for the AIC.
Modeling of FLR data from Narotsky et al. (1997) also gave the same type of
questionable results for the log-logistic model (very low BMDLs relative to the BMD). Here, as
in the case of the Bielmeier et al. (2000) modeling, the initial fit of the log-logistic model does not
appear to be suspect; the goodness of fit evaluations and visual examinations of the model
predictions are consistent with the data and with the other models. It appears that there is some
error or problem with the log-logistic model in the BMDS software that affects the calculation of
BMDLs for some data sets. When the log-logistic model was eliminated from consideration, the
remaining BMDLs are within a factor of 3. The log-probit model was selected because it has the
lowest value for the AIC.
Data from the study by Ruddick et al. (1983) consisted of the count of the numbers of litters that
had one or more fetuses with sternebral variations. Although this expression of the response rates
does not correspond directly to the probability of a response in the offspring of treated dams, it is
consistent with the full litter resorption results from Bielmeier et al. (2001) and Narotsky et al.
(1997) in the sense that it relates to effects recorded at the level of the dam. The log-logistic and
log-probit models could not determine BMDLs for this data set. However, the other models did
provide estimates of the BMDLs, all of which were within a factor of three of one another. The
multistage, having the lowest AIC of all the models was selected as the basis for the BMDL
estimate for this data set.
b. One-day Health Advisory
Four data sets were modeled in support of the One-day HA for bromodichloromethane.
For the Lilly et al. (1994) data on vacuolar degeneration in male rats, the multistage model gave
questionable results (very high AIC and a goodness of fit p value that appeared unrealistically high
Draft -Do not cite or quote A - 35 February 20, 2002
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when the model fit was examined visually) and was eliminated from consideration. The BMDS
software gave warnings on bound calculation for the probit model. All of the remaining BMDLs
are within factor of 3, so the log-probit model was selected because it has the lowest value for the
AIC.
When the data from the same study for renal tubular degeneration were modeled, the
multistage and log-logistic gave questionable results and were eliminated from consideration. The
remaining BMDLs are not within a factor of 3, indicating some model dependence of the results.
The lowest BMDL was thus selected as a reasonable conservative estimate. The gamma or
Weibull models predict the same BMDL. The gamma model was selected on the basis of having
the lowest AIC.
It is perhaps informative to compare the considerations applied here, in the case of the
renal endpoint, to those applied above to the Narotsky et al. (1997) modeling results. In the case
of the Narotsky et al. (1997) results, a single model (log-logistic) gave a BMDL10 that was about
8-fold lower than the corresponding BMD10, whereas the other models gave BMDLs that were
within a factor of about 2 of the corresponding BMDs. The discrepancy was even greater at a
BMR of 5%, suggesting that a problem may be associated with that one model. In the case of the
Lilly et al. (1994) renal effect, after eliminating the obviously problematic model results (log-
logistic and multistage), the differences between BMDs and corresponding BMDLs are present
and consistent for all the models at both 5% and 10% response. It is true that some models have
a greater difference between the BMD and the BMDL than do other models and that this model
dependence is due, at least in part, to the fact that the BMD estimates fall below the lowest
nonzero experimental dose. But the choice of the most conservative BMDL is intended to cover
that model dependence: if there is little information in the region of interest, so that otherwise
reasonable (good-fitting) models disagree as to the BMDL because of differences in possible
curve shapes, the most conservative choice is a good one since we can not rule out the possibility
that the true curve shape is described by the most conservative model. This use of the BMD
methodology and treatment of model-dependence is much superior to the choice of some other
(higher) BMR to use in cases where response rate at the lowest nonzero experimental dose is
greater than 10%. Such an alternative would entail additional arbitrary decisions about what the
higher BMR should be and how to scale the results corresponding to that BMR so as to be
consistent with results from studies in which BMDL10s were estimated.
Renal and hepatic histopathology data were modeled from the study of Thornton-Manning
et al. (1994). The data for hepatic centrilobular vacuolar degeneration were not satisfactorily fit
(all goodness of fit p values were less than 0.1) by any model. These data displayed some
peculiarities: 0% response at the lowest nonzero dose, 100% response at the next dose, and then
a drop to 75% response at the highest dose. Because of the poor fit, the BMDLs for this
endpoint can not be used. For renal data, the multistage model gave questionable results (very
high AIC and a goodness of fit p value that appeared unrealistically high when the model fit was
examined visually) and was eliminated from consideration. The remaining BMDLs were within a
factor of 3, and the log-logistic model was selected because it has the smallest AIC.
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c. Ten-day Health Advisory
Five data sets were modeled using the BMDS software in support of the Ten-day HA for
bromodichloromethane. For the Condie et al. (1983) data on liver histopathology, the logistic and
probit models were eliminated on the basis of poor fit. The rest of the BMDLs were within a
factor of 3, so the log-logistic model was selected on the basis of the smallest AIC. When renal
histopathology data from the same study were modeled, the logistic and probit models were
eliminated for lack of fit. The remaining BMDLs were within a factor of 3, so the models with
the lowest AIC (Weibull and log-logistic) were examined. The Weibull model results were
selected on the basis of the smallest BMDL. For the Aida et al. (1992a) data set for liver cell
vacuolation, questionable results were obtained with the log-logistic model (very low BMDLs
relative to the BMD). The remaining BMDLs are within a factor of 3 and the multistage model
was selected because it had the smallest AIC value. When the Melnick et al. (1998) data were
analyzed, the multistage model was eliminated because it gave a goodness of fit p-value that was
unrealistically high when the curve fit was evaluated by visual inspection and because the AIC was
very large. The BMDL values of the remaining models were within a factor of three. The result
from the Weibull model was selected on the basis of having the lowest AIC value. When results
for the NTP (1998) study were examined, all models gave acceptable fit.- Because all BMDLs
were within a factor of three of one another, the log-logistic model was selected on the basis of
the smallest AIC value.
d. Longer-term Health Advisory
A single data set (NTP, 1987) was analyzed using the BMDS software in support of the
Longer-term HA. Results for the logistic and probit models were rejected for lack of fit
(goodness of fit p values less than 0.1). The remaining p values were within a factor of 3, so the
log-probit model was selected on the basis of the smallest AIC.
e. RfD
Several data sets that had previously given low BMDL10 estimates when modeled using
the Crump benchmark dose software (THC and THWC programs; K.S. Crump, Inc.) were
reanalyzed using the BMDS software and current guidance for evaluation of results. An
advantage of the BMDS package is that it includes several additional model options for data
analysis. Results for the models (Weibull, multistage) common to both programs were in close
agreement. However, one or more of the additional models available in the BMDS sometimes fit
the data better when analyzed by the criteria set forth in Section C (above). In these cases the
BMD and BMDL values changed by a small amount. Where appropriate, these revised values
were used to calculate health advisories and RfDs.
Data for kidney cytomegaly from the chronic NTP (1987) study in male mice were
remodeled using the BMDS program. The results from the logistic and probit models were
rejected for lack of fit (p <0.10). Estimates of the BMDL10 calculated by the Weibull and
multistage models were identical to estimates derived using the Crump benchmark dose software
(0.96 mg/kg-day). The results for the BMDL10 varied by more than a factor of 3, indicating a
Draft -Do not cite or quote A - 37 February 20, 2002
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degree of model dependence. The log-logistic model gave a very low value for the BMDL10 and
was eliminated from further consideration. Of the remaining models, the log-probit model gave
the lowest value for the AIC and the corresponding BMDL10 was thus selected as a candidate for
derivation of the RfD.
When data for fatty degeneration in the liver of female rats (Aida et al., 1992b) were
remodeled using the BMDS program, all models fit the data adequately. The resulting BMDL10
values were within a factor of three, indicating model independence. The BMDL10 calculated
using the probit model was selected as a candidate for derivation of the RfD because it had the
lowest AIC value.
f. Modeling of Continuous Endpoints
Four continuous data sets for maternal body weight gain were modeled in support of
Health Advisory and RfD derivation: Narotsky et al. (1997); CCC (2000b) and CCC (2000d).
Data fitting problems were encountered when attempting to model these data sets using BMDS
Version 1.2. Therefore, these data sets were modeled using BMDS Version 1.3. This version
was not available when the analysis of dichotomous data sets for other endpoints was performed.
For the Narotsky et al. (1997) study, body weight gain data for gestation days 6 to 8 were
modeled. To facilitate modeling, a constant value of 20 was added to each mean so that all
modeled data were positive. This procedure is considered an acceptable approach for
transforming data prior to modeling continuous data with the BMDS software (W. Setzer, U.S.
EPA, personal communication). The BMDS tests for variance rejected the hypothesis that there
is a constant variance for this data set. The modeled variance available in BMDS did a good job
of describing the variation in the variances (p-value of 0.76), when a constant value of 20 was
added to the means.
When the models were fit to the transformed data, modeling the nonconstant variance in
terms of the means plus a constant value of 20, the fits to the means (plus 20) were all good.
Note that the number of parameters for the power and polynomial models are misspecified in
BMDS (because the power hits the bound of 1 and the polynomial is linear), and so the AIC and
p-value for fit are incorrect. The correct values can be found in the output for the linear model.
The results for the Narotsky et al. (1997) body weight gain data are summarized below:
Model
Log-likelihood
AIC
BMD
BMDL
Power (linear)
-84.84
177.68
12.0
9.0
Polynomial
-84.84
177.68
12.0
9.0
Hill
-82.83
177.66
18.3
10.2
Even though the Hill model has two more parameters than the power (linear) model, the
decrease in the log-likelihood gained by those extra parameters is enough to give the slight edge
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February 20, 2002
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to the Hill model in terms of AIC. Thus, the model of choice is the Hill model, since the BMDs
were within a factor of 3 of one another).
To validate the choice of the constant used in the modeling of this data set, we
investigated the effect of adding different constants to the means, with respect to the estimates
from the Hill model. Different constants added to the mean change the parameter estimates
obtained in the maximization of the likelihood. In fact, the choice of the constant can be viewed
as the determination of another parameter that gives the best fit of the model to the data - in this
case allowing the model for the variance to be improved. Note that benchmark responses
(BMRs) defined in terms of a change in the mean equal to some multiple of the control standard
deviation will be appropriate even with the transformed data, because adding a constant to a set of
observations does not alter the standard deviation of the transformed data. Thus, the choice of
the BMR defined as 1.1 standard deviations is consistent for any choice of added constant. (Any
differences in BMDs and BMDLs noted with different added constants is due to differences in the
fitted model parameters, not the definition of the BMR).
In addition to the added constant of 20 that was used for the comparisons above, we also
examined adding constants of 10, 15, 25, or 30. The results of adding the different constants on
the outcome of the Hill model are summarized below:
Constant Added
Log-likelihood
BMD
BMDL
10
-82.83
18.7
11.1
15
-82.73
18.4
10.6
20
-82.71
18.3
10.2
25
-83.37
19.2
10.4
30
-84.22
19.9
8.5
Because the log-likelihood measures the goodness of fit, it can be seen that the constant of
20 is the best choice from among those that we tried. Since the changes in the BMDs and
BMDLs are minor in the region of 20, we did not attempt to fine-tune the choice of the constant
any further. For the Narotsky et al. (1997) data set, the Hill model applied to the data (with a
constant of 20 added to the means) was selected as the best basis for BMD estimation. We
confirmed that the polynomial and power models (which both still defaulted to a linear form) did
not yield a log-likelihood as large as that from the Hill model, for the choice of 20 as the added
constant. The Hill model yielded a BMD of 18, with a BMDL of 10.
For the CCC (2000b) study, data for body weight gain in rabbits on gestation days 6 to 29
(corrected for gravid uterine weight) were modeled. For this data set, the hypothesis of constant
variance could not be rejected at the 0.05 level (p = 0.20). Thus, for all of the modeling
considered, we have assumed constant variance.
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February 20, 2002
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The best-fitting polynomial model was linear. Unfortunately, the linear model did not
describe the data well (p-value for goodness of fit less than 0.001). In contrast, both the power
model and the Hill model gave adequate fits to the data (p-values of 0.28 and 0.52, respectively).
The following table summarizes the outputs for the various models:
Model
Log-likelihood
AIC
BMD
BMDL
Polynomial
(linear)
141.10
-276.2
35.4
29.3
Power
146.47
-284.94
50.3
Failed
Hill
147.04
-284.08
53.7
Failed
Even though the Hill model provided a slightly larger log-likelihood, the gain was not
sufficient to decrease the AIC below that associated with the power model (the Hill model uses 1
extra parameter, and thus the comparison of the AICs says that the improvement in the fit — the
log-likelihood — is not enough to make up for the fact that there is that one extra parameter).
[Note: the AICs in the BMDS output files are incorrectly calculated because the log-likelihoods
are positive numbers rather than negative numbers.] The power model would be the model of
choice, given that the BMDs for the two models that fit the data are within a factor of 3.
However, BMDS did not complete the calculation of the BMDL for either the power or the Hill
model. Nevertheless, it appears that the magnitude of the BMD (around 50 mg/kg) would not
make this endpoint the critical one with respect to finding a health-protective starting point for
RfD determination (i.e., compare 50 mg/kg to the BMDs from other endpoints).
Two data sets for body weight gain were modeled for the CCC (2000d) study in rats. The
decrease in body weight gain was most severe on gestation days 6 to 7. Here, as in the case of
the Narotsky et al. (1997) data set, a constant needed to be added to make the model of the
variance acceptable. However, this data set was problematic. No model available in BMDS fit
the data, regardless of how they were transformed by adding a constant. The power and
polynomial models defaulted to the linear form which clearly did not describe the change in the
means as a function of dose. The Hill model had the required curvature to describe the pattern of
the means as a function of dose, but the problem for that model (as well as for the others) was
that no good model of the variance can be determined just by adding a constant to observations.
The best constant found was around 10, which reduced the log-likelihood for the fitted model to -
246.17. This is compared to the log-likelihood for the independent means, independent variances
model of -240.49. The comparison of these two models would not accept the fitted model (p-
value slightly less than 0.025). As can be seen on the figure in the electronic Appendix B for this
particular choice, the means are well fit by the model, but the variances are not well-modeled,
especially in the control group, for which the estimated standard deviation is greater than the
observed standard deviation. The BMD for that run is 17.7 and the BMDL is 14.5. These are the
smallest from among the runs that we made (using added constants of between 0 and 100, and
also the constant variance model), for which the BMDs ranged between 18.3 and 19, and the
BMDLs ranged between 15.3 and 15.7. It is clear that the BMD and BMDL estimates are not
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A-40
February 20, 2002
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especially sensitive to the choice of the added constant. Considering that the models overpredict
the observed control standard deviation, these BMD estimates may be viewed as slight
overestimates, if anything. But because the fits are not particularly good, some caution might be
warranted if one is considering using these results as the basis for regulation.
To obtain a more reliable estimate of the BMD for decreased body weight observed in
CCC (2000d), body weight gain data were also modeled for gestation days 6 to 9. These data
also required transformation with a constant. This was accomplished by starting the search for
the constant with a minimum of 30 and a maximum of 500. In this case, it was determined that a
value of about 250, added to the means, produced an acceptable model of the change in variance
as a function of the mean (p-value of about 0.21), and yielded the largest maximized likelihood for
the power model. This result was first compared to other choices of constants around 250 (e.g.,
240 and 260), then to values that were progressively further away (e.g., values of 200 and 300, 90
and 400, etc.). We did not fine-tune the estimate of the constant, since the BMD estimates were
stable when the added constant was 240, 250, or 260.
The power model was selected for the above search for the constant because it provided a
much better fit to the transformed data than did the Hill model or the polynomial model for
preliminary choices for the added constant (30 and 55). Because the final choice of the constant
was so different from the preliminary choices, we compared the fits of the models to the
transformed data using 250 as the added constant and the results are summarized here:
Model
Log-likelihood
AIC
BMD
BMDL
Power
-304.942
619.885
22.9
18.4
Polynomial
-336.201
682.403
34.0
Failed
Hill
-413.929
839.858
Failed
Failed
The Hill model did not fit the data at all (perhaps due to a vanishingly small estimate of the
parameter k, which may be due to the fact that the model did not correctly maximize the
likelihood). The polynomial model did not fit the data well and is much less satisfactory than the
power model (compare the AICs in the table above). Even for the power model, the p-value for
goodness of fit was 0.036, which is less than the standard critical p-value of 0.05. However, for
this data set which is nonmonotonic, the power model does a satisfactory job. Consequently, the
reasonable choice for the BMD is 23 and for the BMDL is 18, for this data set.
Table A-5 Benchmark Dose Modeling Results for Bromodichloromethane (Dichotomous
Endpoints)
Draft -Do not cite or quote
A - 41
February 20, 2002
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Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
DEVELOPMENTAL OR REPRODUCTIVE STUDIES
Bielmeier et al. (2001) Rat Female Full Litter Resorption
Gamma
1.0
24.9923
3(2)
36
2.1
42
4.3
Logistic
0.77
25.0528
2(2)
31
8.6
40
16
Log-logistic
1.0
24.9223
3(2)
40
0.8
46
1.6
Log-probit
1.0
24.9223
3(2)
41
5.3
45
7.7
Multistage
0.91
23.1120
1
16
2.1
23
4.2*
Probit
0.83
24.9874
2(2)
28
7.9
37
15
Weibull
1.0
24.9223
3(2)
26
2.1
34
4.3
Narotsky et al. (1997) Rat Female Full Litter Resorption
Gamma
0.68
30.3049
3(2)
36
11
49
22
Logistic
0.47
31.0964
2(2)
41
25
54
39
Log-logistic
0.68
30.3427
3(2)
36
0.13 (?)
48
5.9 (?)
Log-probit
0.72
30.1687
3(2)
36
21
48
30*
Multistage
0.68
30.4257
2
33
10
47
21
Probit
0.53
30.8108
2(2)
40
23
52
37
Weibull
0.67
30.3983
3(2)
35
10
49
22
Ruddick et al. (1983) Rat Female Sternebral Aberrations
Gamma
0.44
64.2926
3(3)
16
7.1
30
15
Logistic
0.66
62.5139
2(2)
22
14
43
28
Log-logistic
0.47
64.2131
3(3)
19
Failed
33
Failed
Log-probit
0.49
64.1729
3(3)
23
Failed
37
Failed
Multistage
0.74
62.2980
2
13
7.1
27
15*
Probit
0.67
62.5023
2(2)
22
14
42
27
Weibull
0.44
64.2965
3(3)
14
7.1
29
15
CANDIDATE STUDIES FOR 1-DAY HA
Lilly et al. (1994) Rat Male Hepatic midzonal vacuolar degeneration (Aqueous vehicle)
Gamma
0.63
3.8029
3(1)
187
135
206
156
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A-42
February 20, 2002
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Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Logistic
1.0
4.0000
2(2)
290
145
292
174
Log-logistic
0.99
2.0468
3(1)
240
164
250
182
Log-probit
1.0
2.0000
3(1)
258
168
263
182*
Multistage
1.0 (?)
298.31
2
6.4
0.1
9.2
0.28
Probit
1.0
4.0000
2(2)
281
184 (W)
285
187 (W)
Weibull
1.0
2.0006
3(1)
294
144
307
170
Lilly et al. (1994) Rat Male Renal Tubule degeneration (Aqueous vehicle)
Gamma
1.0
9.6389
3(1)
119
4.3
131
8.9*
Logistic
1.0
11.6382
2(2)
163
19
171
35
Log-logistic
1.0
9.6382
3(1)
163
0.00078
(?)
170
0.00625
(?)
Log-probit
1.0
11.6382
3(2)
152
11
160
16
Multistage
1.0 (?)
102.1000
2
6.4
0.13
9.2
0.12
Probit
1.0
11.6302
2(2)
132
17
144
33
Weibull
1.0
11.6302
3(2)
92
4.4
110
8.9
Thornton-Manning et al. (1994) Rat Female Hepatic centrilobular vacuolar degeneration (Poor fit: no model
selected
Gamma
0.01
19.0705
3(2)
40
5.5
53
11
Logistic
<0.01
21.4063
2(2)
34
16
54
29
Log-logistic
<0.01
17.3420
3(2)
55
13
67
22
Log-probit
0.02
17.9809
3(2)
53
16
64
22
Multistage
0.01
20.1038
2
17
5.0
31
10
Probit
<0.01
219.846
2(2)
31
15
52
29
Weibull
0.02
19.6598
3(2)
23
5.1
37
10
Thornton-Manning et al. (1994) Rat Female Renal tubular degeneration
Gamma
1.0
10.3742
3(1)
99
45
109
60
Logistic
1
12.3178
2(2)
138
42
141
63
Log-logistic
1
10.3178
3(1)
127
51
133
65*
Log-probit
1
12.3178
3(2)
123
54
128
66
Draft -Do not cite or quote
A-43
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Multistage
1 (?)
445.4650
2
3.4
0.06
4.4
0.08
Probit
1
12.3178
2(2)
128
38
133
58
Weibull
1
12.3179
3(2)
127
39
133
56
CANDIDATE STUDIES FOR THE 10-DAY HA
Condie et al. (1983) Mouse Male Hepatic centrilobular pallor
Gamma
0.19
30.0268
3(2)
11
2.2
17
4.5
Logistic
<0.01
33.9633
2(2)
14
7.9
23
14
Log-logistic
0.30
28.7294
3(2)
19
4.4
24
7.5*
Log-probit
0.26
29.2521
3(2)
18
6.0
23
8.6
Multistage
0.16
30.6962
2
4.2
2.1
8.5
4.3
Probit
<0.01
34.7900
2(2)
13
7.6
22
14
Weibull
0.18
30.3429
3(2)
7.2
2.1
12
4.4
Condie et al. (1983) Mouse Male Renal epithelial hyperplasia
Gamma
0.88
35.6018
3(2)
75
46
83
56
Logistic
0.10
40.0819
2(2)
25
15
40
27
Log-logistic
0.98
35.3785
3(2)
112
48
117
58
Log-probit
0.84
37.3705
3(3)
109
51
113
59
Multistage
0.38
37.5968
2
46
17
58
31
Probit
0.08
40.7541
2(2)
21
13
35
24
Weibull
0.98
35.3785
3(2)
120
41
125
53*
Aida et al. (1992a) Rat Female Liver cell vacuolization
Gamma
0.56
23.4101
3(2)
18
8.5
36
17
Logistic
0.23
25.2544
2(2)
50
27
83
49
Log-logistic
0.60
23.2879
3(2)
19
0.0015 (?)
35
0.1 (?)
Log-probit
0.65
21.5114
3(2)
34
20
49
28
Multistage
0.78
21.4146
1
16
8.5
34
17*
Probit
0.25
25.0481
2(2)
46
25
76
46
Weibull
0.57
23.4143
3(2)
17
8.5
34
17
Draft -Do not cite or quote
A-44
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Melnick et al. (1998) Mouse Female Hepatocyte hydropic degeneration
Gamma
0.99
26.2625
3(2)
28
4.2
35
8.5
Logistic
0.82
26.8455
2(2)
24
11
36
20
Log-logistic
0.91
26.5313
3(2)
31
11
38
16
Log-probit
0.96
26.3532
3(2)
32
12
38
17
Multistage
1.0?
445.4650
2
3.4
0.076
4.4
0.082
Probit
0.89
26.5905
3(2)
24
11
34
19
Weibull
1.0
26.2278
3(2)
23
4.2
31
8.4*
NTP (1998) Rat Male Single cell hepatic necrosis
Gamma
1.0
14.4941
3(1)
26
12
28
15.125
Logistic
1.0
16.3653
2(2)
34
12
35
16.9848
Log-logistic
1.0
14.3662
3(1)
33
15
34
18.4508*
Log-probit
1.0
16.3653
3(2)
33
15
334
17.4679
Multistage
0.93
15.1127
1
16
4.8
20
9.4
Probit
1.000
16.3653
2(2)
31
10
32
15
Weibull
1.000
16.3653
3(2)
30
9.8
32
14
CANDIDATE STUDIES FOR THE LONGER-TERM HA
NTP (1987) Mouse Female Hepatic Vacuolated Cytoplasm
Gamma
0.27
32.2103
3(2)
57
28
73
42
Logistic
0.04
37.1630
2(2)
54
33
80
54
Log-logistic
0.38
31.1109
3(2)
60
33
74
46
Log-probit
0.41
30.9082
3(2)
62
35
75
47*
Multistage
0.32
31.6422
1
44
20
63
37
Probit
0.05
36.4925
2(2)
55
32
80
54
Weibull
0.20
33.6311
3(2)
46
21
65
35
CANDIDATE STUDIES FOR THE RfD
Aida et al. (1992b) Rat Female Hepatic Fatty Degeneration
Gamma
1.0
46.8266
--
5.1
0.71
5.8
1.4
Draft -Do not cite or quote
A-45
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Logistic
0.99
44.8652
--
2.0
1.3
3.4
2.3
Log-logistic
1.0
42.8266
--
6.7
2.0
7.1
2.9
Log-probit
1.0
46.8266
--
5.9
2.0
6.4
2.7
Multistage
1.0
48.8266
--
2.9
0.56
4.0
1.1
Probit
1.0
44.8296
--
1.8
1.2
3.1
2.1*
Weibull
1.0
46.8266
--
3.3
0.65
4.4
1.3
Aida et al. (1992b) Rat Male Hepatic Fatty Degeneration
Gamma
1.0
29.3001
3(2)
1.2
0.39
2.1
0.80
Logistic
0.15
34.2633
2(2)
2.6
1.5
4.4
2.7
Log-logistic
0.98
29.3760
3(2)
2.1
0.51
3.0
0.94
Log-probit
1.0
29.3074
3(2)
2.2
0.99
3.0
1.4
Multistage
1.0
29.3001
2
0.73
0.43
1.5
0.88
Probit
0.17
33.8989
2(2)
2.6
1.6
4.4
2.9
Weibull
1.0
29.3001
3(2)
1.1
0.39
1.9
0.80*
Aida et al. (1992b) Rat Male Hepatic Granulomas
Gamma
0.99
34.9241
3(1)
1.0
0.67
2.1
1.4
Logistic
0.13
42.0098
2(2)
4.2
2.6
7.0
4.6
Log-logistic
0.66
38.0328
3(2)
1.9
0.37
3.0
0.82
Log-probit
0.87
35.6192
3(1)
2.4
1.6
3.5
2.3
Multistage
0.95
36.8960
2
1.1
0.67
2.2
1.4
Probit
0.15
41.5172
2(2)
3.8
2.4
6.4
4.4
Weibull
0.99
34.9241
3(1)
1.0
0.67
2.1
1.4*
NTP (1987) Mouse Male Hepatic Focal Necrosis
Gamma
1.0
15.6352
3(1)
44
27
49
34
Logistic
1.0
17.4602
2(2)
65
30
66
38
Log-logistic
1.0
15.4604
3(1)
59
28
61
34
Log-probit
1.0
17.4602
3(2)
57
28
60
34
Multistage
1.0
16.0769
1
40
23
47
32
Draft -Do not cite or quote
A-46
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Probit
1.0
17.4602
2(2)
60
28
62
36
Weibull
1.0
15.4603
3(1)
60
27
63
35*
NTP (1987) Mouse Male Kidney Cytomegaly
Gamma
0.76
72.3705
0.58
0.47
1.2
0.96
Logistic
<0.01
88.5494
3.3
2.1
5.3
3.8
Log-logistic
1.0
73.8361
1.5
3.6E-9
2.2
1.1E-7
Log-probit
0.99
71.8572
1.4
1.1
2.0
1.5*
Multistage
0.45
74.3705
0.58
0.47
1.2
0.96
Probit
<0.01
92.6760
2.8
2.0
4.8
3.6
Weibull
0.76
72.3705
0.58
0.47
1.2
0.96
* Selected model result for endpoint.
AIC Akaike Information Criterion
BMD Benchmark Dose
BMDL 95% lower confidence level on BMD
G-O-F Goodness-of-Fit
HA Health Advisory
? Results questionable on the basis of visual inspection or probable calculation error
W BMDS gave a warning message: "BMDL computation is at best imprecise for these data"
Draft -Do not cite or quote
A-47
February 20, 2002
-------
2. Dibromochloromethane
BMDS modeling results for dibromochloromethane are summarized in Table A-6 below.
Detailed output for each model run is compiled in Appendix B, provided in electronic format on
compact disk.
a. Developmental and Reproductive Studies
No data were modeled. The generation F2b day 14 postnatal body weight data of
Borzelleca and Carchman (1982) were considered for modeling. However, the study authors did
not report the number of litters examined for this continuous endpoint. Since this information is
required as input, the data could not be modeled.
b. One-day Health Advisory
No data were modeled for the One-day HA.
c. Ten-day Health Advisory
Seven data sets were modeled in support of the Ten-day HA. When data were analyzed
for the liver cell vacuolation in female rats (Aidal992a), the multistage model gave questionable
results and was eliminated from consideration. The remaining BMDL values were within a factor
of 3, so the estimate from the Weibull and gamma models was selected on the basis of the smallest
AIC. For the same endpoint in male rats (Aida et al. 1992a), model fit was adequate in all cases
and all BMDL values were within a factor of 3. The multistage model was selected on the basis
of the smallest AIC.
Liver and kidney histopathology data from the study by Condie et al. (1983) were
analyzed. In the analysis of kidney data, the BMDLs calculated by the logistic and probit models
were eliminated because the models fit the data poorly. Among the remaining models, the log-
logistic BMDLs are smallest by more than a factor of 3. Thus, this result was examined as a
possible outlier. Although this model does at times have difficulty calculating reasonable lower
bounds, the BMDs in this case are also smaller than the BMDs from the other models (although
not by a factor of 3 for all them - the BMDLs calculated by the logistic model are more than 3-
fold below any other BMDL). In addition, the AIC value is much lower for log-logistic than for
the other models. Thus, the BMDL calculated by the log-logistic model was selected.
With respect to the Condie et al. (1983) liver histopathology data, the log-logistic model
gives the smallest BMDL by more than a factor of 3. However, the BMDs are within a factor of
3 of the BMDs generated by other models and the AIC for the log-logistic model is not the
lowest. The log-logistic results were therefore considered as outliers. Among the remaining
options, the Weibull and gamma models have the lowest AIC, and the BMDL calculated by these
models was selected.
Draft -Do not cite or quote
A-48
February 20, 2002
-------
Data for stomach nodules in male and female rats (NTP, 1985) were also analyzed. In
females, all calculated BMDLs are within a factor of 3, so the result from the multistage model
was selected on the basis of having the smallest AIC. In males, the log-logistic model has the
smallest AIC and the lowest BMDL.
In the analysis of the Melnick et al. (1998) results for hepatic hydropic degeneration in
female mice, the multistage model failed to fit the data. All remaining results were similar and the
BMDL calculated by the log-logistic model was selected on the basis of the lowest AIC.
d. Longer-term Health Advisory
Two data sets for hepatic lesions reported in Chu et al. (1982b) were modeled in support
of the Longer-term HA. When liver histopathology data for male rats was analyzed, the log-
logistic model calculated a BMDL that was more than 3-fold lower than those from some other
models. However, both the AIC and the BMD calculated by the log-logistic model were the
lowest among all models. The BMDL calculated by this model was thus selected. When data for
liver lesions in female rats were analyzed, no model adequately fit the data (all p values were less
than 0.1). Thus, no BMDL was selected from this data set.
Data for fatty metamorphosis in the liver of male rats that had previously been modeled
using the Crump software was reanalyzed using the BMDS program. All models adequately fit
the data for this endpoint. All resulting BMDL values were within a factor of three, with the
exception of the estimate calculated using the log-probit model The value calculated using the
probit model was selected on the basis of the lowest AIC.
e. RfD
Two data sets from the NTP (1985) chronic oral exposure study were modeled using the
BMDS software in support of the RfD for dibromochloromethane. These data sets were selected
after inspection of the results for BMD modeling of key dibromochloromethane endpoints using
the Crump software (K. S. Crump, Inc.). For fatty metamorphosis in the liver of male rats, all
models fit the data acceptably, although the p value for the probit model was marginal (p = 0.16).
All BMDL values were within a factor of three, with the exception of the log-logistic model.
Results from the log-logistic model were selected on the basis of the lowest AIC value. For
ground glass cytoplasm in the liver of male rats, all models fit the data acceptably and the BMDL
values were within a factor of three. The results for the probit model were selected on the basis
of the lowest AIC.
Draft -Do not cite or quote
A-49
February 20, 2002
-------
Table A-6 Benchmark Dose Modeling Results for Dibromochloromethane
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
DEVELOPMENTAL AND REPRODUCTIVE STUDIES (NONE)
CANDIDATE STUDIES FOR 1-DAY HA (NONE)
CANDIDATE STUDIES FOR 10-DAY HA
Aida et al. (1992a) Rat Female Liver cell vacuolization
Gamma
1.0
15.4833
3(2)
24
3.2
30
6.7
Logistic
0.91
15.7862
2(2)
24
10
34
17
Log-logistic
0.99
15.5265
3(2)
24
7.5
30
12
Log-probit
1.0
15.4078
3(2)
25
8.4
30
12
Multistage
1 (?)
347
2
3.4
0.063
4.4
0.064
Probit
0.95
15.6586
2(2)
22
8.9
32
16
Weibull
1.0
15.4833
3(2)
21
3.2
29
6.7*
Aida et al. (1992a) Rat Males Liver cell vacuolization
Gamma
0.76
20.0153
3(2)
12
2.5
18
5.1
Logistic
0.77
20.0459
2(2)
17
8.0
26
14
Log-logistic
0.56
20.8287
3(2)
14
2.7
20
5.3
Log-probit
0.57
20.7294
3(2)
14
6.0
19
8.6
Multistage
0.98
19.3422
2
7.0
2.7
14
5.5*
Probit
0.80
19.9157
2(2)
15
7.4
24
13
Weibull
0.83
19.7280
3(2)
12
2.6
18
5.3
Condie et al (1983) Mouse Male Renal mesangial hypertrophy
Gamma
0.12
36.8675
3(1)
3.8
2.6
7.8
5.3
Logistic
<0.01
47.0265
2(2)
12
7.7
22
15
Log-logistic
0.49
33.8237
3(1)
16
0.7
3.5
1.6*
Log-probit
0.11
36.7799
3(1)
8.9
5.6
13
8.1
Multistage
0.12
36.8675
2(1)
3.8
2.6
7.8
5.3
Probit
<0.01
46.9299
2(2)
12
8.0
22
16
Weibull
0.12
36.8675
3(1)
3.8
2.6
7.8
5.3
Draft -Do not cite or quote
A-50
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Condie et al. (1983) Mouse Male hepatic cytoplasmic vacuolization
Gamma
0.32
43.8699
3(2)
5.4
3.4
11
6.9
Logistic
0.15
45.1424
2(2)
15
9.4
27
18
Log-logistic
0.33
43.9124
3(2)
3.3
1.5
7.0
3.3
Log-probit
0.23
44.4103
3(2)
14
8.5
20
12
Multistage
0.13
45.8578
3
5.7
3.4
12
6.9
Probit
0.16
44.9950
2(2)
14
9.2
26
18
Weibull
0.32
43.8699
3(2)
5.4
3.4
11
6.9*
NTP (1985) Mouse Female Stomach nodules
Gamma
0.98
16.2743
3(2)
167
38
225
78
Logistic
0.86
17.1462
2(2)
209
104
284
170
Log-logistic
0.98
16.2955
3(2)
163
34
222
73
Log-probit
0.99
16.1399
3(2)
166
74
218
106
Multistage
0.99
14.3879
1
152
37
218
77*
Probit
0.90
16.8628
2(2)
197
95
267
158
Weibull
0.97
16.3761
3(2)
162
37
227
77
NTP (1985) Mouse Male Stomach nodules
Gamma
0.91
18.5058
3(1)
75
33
153
67
Logistic
0.64
21.8294
2(2)
191
97
306
168
Log-logistic
0.93
18.5021
3(1)
68
26
143
54*
Log-probit
0.68
19.4319
3(1)
122
69
176
99
Multistage
0.91
18.5858
1
75
33
153
67
Probit
0.65
21.7110
2(2)
174
88
284
154
Weibull
0.91
18.5858
3(1)
75
33
153
67
Melnick et al (1998) Mouse Female hepatic hydropic degeneration
Gamma
0.99
12.6487
3(1)
76
55
84
64
Logistic
1.0
14.0080
2(2)
123
56
126
70
Log-logistic
1.0
12.0080
3(1)
108
59
112
68*
Draft -Do not cite or quote
A - 51
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Log-probit
1.0
14.0080
3(2)
107
60
111
68
Multistage
program
failed!
--
--
--
--
--
Probit
1.0
14.0080
2(2)
112
56
116
68
Weibull
1.0
14.0080
3(2)
112
56
117
68
CANDIDATE STUDIES FOR LONGER-TERM HA
Chu et al. (1982b) Rat Male Hepatic lesions
Gamma
0.84
65.3956
3(2)
14
6.2
29
13
Logistic
0.79
65.6217
2(2)
22
12
43
23
Log-logistic
0.88
65.1734
3(2)
8.6
2.5
18
5.3*
Log-probit
0.72
65.8723
3(2)
37
15
54
22
Multistage
0.84
65.3956
2
14
6.2
29
13
Probit
0.79
65.6176
2(2)
22
12
43
24
Weibull
0.84
65.3956
3(2)
14
6.2
29
13
Chu et al. (1982b) Rat Female Hepatic lesions
Gamma
0.04
67.0864
3(3)
Flat Curve
Estimated
No BMDs
--
Logistic
0.09
64.3869
2(2)
66
26
127
49
Log-logistic
0.09
64.3474
3(2)
48
9.8
101
21
Log-probit
0.04
67.0864
3(3)
4800
38
6900
54
Multistage
0.09
64.3615
2
54
15
110
30
Probit
0.09
64.3843
2(2)
64
25
125
48
Weibull
0.09
64.3615
3(2)
54
15
110
30
NTP 1985 Rat Male Fatty Metamorphosis (Subchronic)
Gamma
0.90
42.3900
3(3)
2.6
0.44
3.9
0.91
Logistic
0.97
40.3442
2(2)
1.2
0.76
2.4
1.5
Log-logistic
0.81
42.9172
3(3)
4.1
0.20
5.5
0.42
Log-probit
0.85
42.6546
3(3)
4.3
1.1
5.4
1.6
Multistage
0.92
43.8670
4
1.0
0.49
2.1
1.0
Probit
0.98
40.1651
2(2)
1.3
0.84
2.5
1.7*
Draft -Do not cite or quote
A-52
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Weibull
0.92
42.2885
3(3)
2.4
0.45
3.9
0.92
CANDIDATE STUDIES FOR THE RfD
NTP (1985) Rat Male Hepatic Fatty Metamorphosis - Chronic
Gamma
0.53
105.8690
3(2)
0.81
0.57
1.7
1.2
Logistic
0.34
106.2850
2(2)
1.2
0.86
2.4
1.7
Log-logistic
1.0
107.4950
3(3)
1.7
0.071
2.7
0.15
Log-probit
0.86
105.5270
3(2)
1.9
1.1
2.7
1.6*
Multistage
0.53
105.8690
2
0.81
0.57
1.7
1.2
Probit
0.16
107.2230
2(2)
1.4
1.1
2.8
2.2
Weibull
0.53
105.8690
3(2)
0.81
0.57
1.7
1.2
NTP (1985) Rat Male Ground Glass Cytoplasm - Chronic
Gamma
1.0
181.2470
3(3)
6.1
2.4
10
5.0
Logistic
0.58
179.5560
2(2)
6.3
5.1
12
9.7
Log-logistic
1.0
181.2470
3(3)
7.6
1.6
12
3.5
Log-probit
1.0
181.2470
3(3)
9.1
6.3
13
9.1
Multistage
1.0
181.2470
3
4.3
2.4
8.6
5.0
Probit
0.62
179.4870
2(2)
6.0
4.9
11
9.4*
Weibull
1.0
181.2470
3(3)
5.5
2.4
9.8
5.0
* Selected model result for endpoint.
AIC Akaike Information Criterion
BMD Benchmark Dose
BMDL 95% lower confidence level on BMD
G-O-F Goodness-of-Fit
HA Health Advisory
? Results questionable on the basis of visual inspection or probable calculation error
W BMDS gave a warning message: "BMDL computation is at best imprecise for these data"
Draft -Do not cite or quote
A-53
February 20, 2002
-------
3. Bromoform
BMDS modeling results for dibromochloromethane are summarized in Table A-7 below.
Detailed output for each model run is compiled in Appendix B, provided in electronic format on
compact disk.
a. Developmental and Reproductive Studies
Data from the study by Ruddick et al. (1983) consisted of the count of the numbers of
litters that had one or more fetuses with sternebral variations. This expression of the response
rates does not correspond directly to the probability of a response in the offspring of treated
dams. All of the model fit the data and all of the BMDL results for 10% extra risk are within
a factor of three of one another, so the results from the log-probit model were selected, because
that model had the lowest AIC.
b. One-day Health Advisory
BMD calculations were not conducted in support of the One-day HA due to a lack of
appropriate data.
c. Ten-day Health Advisory
Six data sets were modeled in support of the Ten-day HA. Modeling results for each data
set were evaluated using the criteria given in Section C. For the Aida et al. (1992a) data on liver
cell vacuolation in female rats, all models fit well and (with one exception) give the same AIC.
The log-logistic results appear to be qualitative outliers because they are more than 3 times less
than the next closest BMDLs, even though the BMD is the second largest. The BMDL calculated
by this model was thus rejected in favor of the next lowest BMDL (Weibull model). When the
same endpoint was modeled in male rats (Aida et al. 1992), the probit model either failed or gave
a warning message for the lower bound calculations and results were thus eliminated. The
remaining models calculated very similar BMDLs. The Weibull model was selected because it
gave the lowest AIC among the remaining models.
For histopathological effects in the kidney of male mice (Condie et al., 1983), two
models (logistic and probit) had somewhat higher BMDLs and the largest values for AIC. If
these results are eliminated as qualitative outliers, the remaining BMDLs are within a factor of 3.
The Log-probit model was selected from among the remaining models because it gave the lowest
AIC. Modeling of data on liver histopathology from the same study gave a similar pattern of
results. Results from the probit and logistic models were eliminates as qualitative outliers (high
AICs and BMDLs that were higher by more than a factor of 3 from the lowest BMDL). The
remaining BMDLs are within a factor of 3 and so the Log-probit model was selected because it
had the lowest AIC.
Melnick et al. (1998) reported data for hydropic degeneration in the liver of female mice.
The multistage model gave questionable results (very high AIC and a goodness of fit p value that
Draft -Do not cite or quote A - 54 February 20, 2002
-------
appeared unrealistically high when the model fit was examined visually) for this data set that
appeared to reflect a calculation error in the BMDS software. The BMDLs estimated by the
remaining models are very close. The Log-probit model was selected because it has the lowest
AIC. When data for stomach nodules in male mice were modeled (NTP, 1989a), all models gave
an acceptable fit and all BMDLS were within a factor of three. The results from the multistage
model were selected because it had the lowest AIC.
d. Longer-term Health Advisory
Three data sets were modeled in support of the longer-term Health Advisory for
Bromoform. No models adequately fit (i.e. all p values for goodness of fit were less than 0.1) the
Chu et al. (1982b) data for hepatic lesions in female rats (nonmonotonic dose response). Thus,
none of the calculated BMDLs were candidates for deriving the Longer-term Health Advisory.
For the same endpoint in male rats (Chu et al. 1982b), the multistage model gave a bad fit and
was eliminated from consideration. The log-logistic BMDLs are the lowest of the remaining
values, but there is a spread of greater than 3. The result for the log-logistic model was
eliminated as a qualitative outlier, since this model gave the largest BMDs but the BMDLs were
among the lowest observed (i.e. gave a wide confidence interval). The remaining BMDL values
were similar and the probit model was selected because it gave the lowest AIC value. Modeling
of the NTP (1989a) data for hepatic vacuolation in female mice gave similar results across all
models. The Log-probit model was selected because it gave the lowest AIC .
e. RfD
Two data sets from the oral exposure study conducted in rats by NTP (1989a) were
modeled using the BMDS program for consideration in derivation of the RfD. For hepatic
vacuolization in male rats exposed to bromoform for 13 weeks, no fit was obtained for the
multistage model. The BMDL values calculated using the remaining models were with a factor of
3, with the exception of the log-logistic model. The results from the Weibull model were selected
on the basis of the lowest AIC value. With respect to data for fatty changes in the liver of male
rats chronically exposed to bromoform (NTP, 1989a), all models gave acceptable fits. BMDLs
calculated by all models except the log-logistic were within a factor of 3. The log-logistic model
was eliminated as a qualitative outlier, since it gave the highest BMDs but very low BMDLs (i.e.
it resulted in a very wide confidence interval). Of the remaining models, the lowest AIC was
observed for the multistage and it was therefore selected.
Draft -Do not cite or quote
A-55
February 20, 2002
-------
Table A-7 Benchmark Dose Modeling Results for Bromoform
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
REPRODUCTIVE STUDIES
Ruddick et al. (1983) Rat Females Sternebral Aberrations
Gamma
0.85
67.302
3(3)
16
9.2
32
19
Logistic
0.70
66.0887
2(2)
33
23
59
42
Log-logistic
0.90
67.3517
3(3)
19
6.4
35
14
Log-probit
0.87
65.6053
3(2)
35
23
50
33*
Multistage
0.85
67.339
3
15
9.1
31
19
Probit
0.74
65.9551
2(2)
30
21
55
40
Weibull
0.85
67.3711
3(3)
16
9.2
32
19
CANDIDATE STUDIES FOR 1-DAY HA (NONE)
CANDIDATE STUDIES FOR 10-DAY HA
Aida et al. (1992a) Rat Females Liver cell vacuolization
Gamma
1.0
12.3758
3(2)
23
1.1
28
2.3
Logistic
1.0
12.3758
2(2)
45
5.2
47
9.6
Log-logistic
1.0
12.3758
3(2)
42
0.29
45
0.61
Log-probit
1.0
12.3758
3(2)
32
2.6
35
3.7
Multistage
1.0
14.3758
3
9.1
1.9
14
2.4
Probit
1.0
12.3758
2(2)
36
4.8
40
9.0
Weibull
1.0
12.3758
3(2)
11
1.1
16
2.3*
Aida et al. (1992a) Rat Males Liver cell vacuolization
Gamma
0.99
2.2426
3(1)
73
44
81
53
Logistic
1.0
4.0000
2(2)
116
56
118
57
Log-logistic
1.0
2.0014
3(1)
91
48
95
56
Log-probit
1.0
4.0000
3(2)
95
49
93
56
Multistage
0.76
4.2093
1
46
14
59
28
Probit
1.0
4.0000
2(2)
118
failed
121
58 (W)
Weibull
1.0
2.0000
3(1)
134
40
140
51*
Draft -Do not cite or quote
A-56
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Condie et al. (1983) Mouse Male Renal mesangial nephrosis
Gamma
0.46
32.2294
3(2)
45
9.1
64
19
Logistic
0.13
38.0766
2(2)
54
30
85
54
Log-logistic
0.51
31.9498
3(2)
49
6.6
68
14
Log-probit
0.53
31.8168
3(2)
57
23
73
34*
Multistage
0.40
32.6120
2
29
8.7
51
18
Probit
0.16
34.4823
2(2)
52
29
82
52
Weibull
0.44
32.4190
3(2)
35
8.9
56
18
Condie et al. (1983) Mouse Male Centrilobular pallor
Gamma
0.38
30.0854
3(2)
44
7.4
61
15
Logistic
0.10
33.0715
2(2)
46
25
73
45
Log-logistic
0.46
29.5813
3(2)
51
8.8
67
17
Log-probit
0.46
29.5362
3(2)
56
20
70
28*
Multistage
0.31
30.6200
2
29
7.0
49
14
Probit
0.11
32.6482
2(2)
45
25
71
44
Weibull
0.35
30.4002
3(2)
32
7.2
50
15
Melnick et al. (1998) Mouse Female Liver hydropic degeneration
Gamma
0.99
8.3184
3(1)
177
111
196
135
Logistic
1.0
10.2790
2(2)
190
85
199
123
Log-logistic
1.0
8.2790
3(1)
191
123
199
146*
Log-probit
1.0
10.2790
3(2)
189
127
198
146
Multistage
1.0 (?)
396.414
2
64
0.099
9.2
0.16
Probit
1.0
10.2790
2(2)
182
78
197
115
Weibull
1.0
10.2790
3(2)
172
88
196
118
NTP (1989a) Mouse Male Stomach nodules
Gamma
0.43
20.0036
3(2)
165
46
208
82
Logistic
0.22
22.0653
2(2)
158
77
223
131
Log-logistic
0.48
19.6930
3(2)
165
55
206
89
Draft -Do not cite or quote
A-57
February 20, 2002
-------
Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Log-probit
0.50
19.4922
3(2)
176
66
211
95
Multistage
055
18.7358
1
116
32
167
66*
Probit
0.25
21.5725
2(2)
162
74
222
127
Weibull
0.38
20.6462
3(2)
137
35
188
68
CANDIDATE STUDIES FOR THE LONGER-TERM HA
Chu et al. (1982b) Rat Female Hepatic lesions
Gamma
flat
curve
fit
no BMD
--
--
--
Logistic
0.06
53.0088
2(2)
38
23
69
44
Log-logistic
0.07
51.4919
3(2)
10
4.1
21
8.6
Log-probit
0.05
52.6957
3(2)
36
18
51
26
Multistage
0.07
51.9025
2
16
8.2
32
17
Probit
0.06
52.9040
2(2)
35
22
65
42
Weibull
0.07
51.9025
3(2)
16
8.2
32
17
Chu et al.( 1982b) Rat Male Liver lesions
Gamma
0.45
56.7134
3(3)
31
1.6
36
3.3
Logistic
0.66
54.7588
2(2)
5.1
2.8
10
5.6
Log-logistic
0.45
56.7134
3(3)
45
0.8
48
1.6
Log-probit
0.45
56.7134
3(3)
37
4.3
41
6.1
Multistage
<0.01
202.7
2
2.5
1.1
3
1.9
Probit
0.67
54.6647
2(2)
5.3
3.0
10
5.9*
Weibull
0.45
56.7024
3(3)
13
1.6
19
3.3
NTP (1989a) Mouse Female Hepatic vacuolization
Gamma
0.73
30.3981
3(2)
69
35
82
51
Logistic
0.27
33.9813
2(2)
68
41
96
66
Log-logistic
0.80
30.0241
3(2)
70
38
87
54
Log-probit
0.85
29.6545
3(2)
74
41
88
55*
Multistage
0.62
31.6064
2
53
24
76
45
Probit
0.33
33.0847
2(2)
68
40
95
64
Draft -Do not cite or quote
A-58
February 20, 2002
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Model
G-O-F
p value
AIC
No.
paramet.
05
10
BMD
BMDL
BMD
BMDL
Weibull
0.61
31.3715
3(2)
56
28
81
46
CANDIDATE STUDIES FOR THE RfD
NTP (1989a) Rat Male Hepatic vacuolation - Subchronic
Gamma
0.70
65.8698
3(2)
2.2
1.3
4.4
2.6
Logistic
0.65
66.1399
2(2)
3.9
2.4
7.3
4.7
Log-logistic
0.49
68.4402
3(3)
2.5
0.45
4.4
0.94
Log-probit
0.61
66.4811
3(2)
5.5
2.9
8.0
4.2
Multistage
-
661.889
6
2.1
0.0064
2.4
0.0084
Probit
0.65
66.0888
2(2)
3.9
2.6
7.8
5.3
Weibull
0.70
65.8698
3(2)
2.2
1.3
4.4
2.6*
NTP (1989a) Rat Male Fatty Changes in Liver - Chronic
Gamma
1.0
84.7983
3(3)
29
0.66
32
1.4
Logistic
0.89
82.8323
2(2)
1.9
1.2
3.8
2.4
Log-logistic
1.0
82.7983
3(2)
51
0.016
53
0.033
Log-probit
1.0
84.7983
3(3)
38
0.97
41
1.4
Multistage
0.99
82.7983
2
8.9
0.66
13
1.4*
Probit
0.98
82.7996
2(2)
2.2
1.6
4.5
3.2
Weibull
1.0
84.7983
3(3)
12
0.66
16
1.4
* Selected model results for endpoint.
AIC Akaike Information Criterion
BMD Benchmark Dose
BMDL 95% lower confidence level on BMD
G-O-F Goodness-of-Fit
HA Health Advisory
? Results questionable on the basis of visual inspection or probable calculation error
W BMDS gave a warning message: "BMDL computation is at best imprecise for these data"
Draft -Do not cite or quote
A-59
February 20, 2002
-------
APPENDIX B
Appendix B contains BMD Modeling Output in Electronic Format (compact disk)
Draft -Do not cite or quote
B - 1
February 20, 2002
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APPENDIX C
Determination of the Relative Source Contribution for Dibromochloromethane (DBCM)
Draft -Do not cite or quote
C- 1
February 20, 2002
-------
The relative source contribution (RSC) is the percentage of total daily exposure that is
attributable to tap water when all potential sources are considered (e.g., air, food, soil, and
water). Ideally, the RSC is determined quantitatively using nationwide, central tendency and/or
high-end estimates of exposure from each relevant medium. In the absence of such data, a default
RSC ranging from 20% to 80% may be used. The RSC used in the current and previous drinking
water regulations for DBCM is 80%. This value was determined by use of a screening level
approach to estimate and compare exposure from various sources. Information considered for
DBCM during this process is summarized below.
The initial step in RSC determination is problem formulation, including identification of
population(s) of concern, critical health effects, and relevant exposure sources and pathways. The
occurrence of DBCM in tap water is reasonably well documented. Occurrence is widespread as a
result of disinfection of drinking water, resulting in broad exposure of the U.S. general
population. For chronic exposure to DBCM, the most sensitive responses in animal studies are
histopathological changes in the liver. There is no evidence that children or the fetus are more
sensitive to these effects than are adults. Although polymorphisms in metabolizing enzymes might
predispose some groups to greater sensitivity to this compound, no sensitive subpopulations have
yet been clearly identified. Therefore, the population of concern for exposure to DBCM is
considered to be the U.S. general population.
Production and use of DBCM occur mainly on a limited scale in the United States. In the
past, brominated trihalomethanes have been used in pharmaceutical manufacturing and chemical
synthesis, as ingredients in fire-resistant chemicals and gauge fluids, and as solvents for waxes,
greases, resins, and oils (U.S. EPA, 1975). However, use patterns have changed over time.
DBCM is now reportedly used in laboratory quantities only (ATSDR, 1990). Thus, releases to
the environment are not anticipated to be significant on a nationwide basis when compared to
occurrence in disinfected tap water.
DBCM has been detected in air and food in a few studies, in addition to its presence in tap
water. No data were available in the materials reviewed for levels of DBCM in soil. DBCM is
expected to volatilize readily from wet or dry soil surfaces based on its Henry's Law constant and
vapor pressure (U.S. EPA, 1987). For this reason, exposure via ingestion of soil is not expected
to be a significant route of exposure. Therefore, water, food, and air are considered to be the
relevant pathways for this analysis.
Evaluation of Occurrence Data
The next step in RSC determination is to judge whether or not adequate data exist to
characterize exposure from relevant exposure pathways. Factors to be considered in the
evaluation of data adequacy include sample size; whether the data represent a random sample and
are representative of the target population; acceptable analytical detection limits; statistical
distribution of the data, and estimator precision. In addition, it is important to know whether the
data are representative of current conditions. The available occurrence data for DBCM in water,
air, food, and soil are summarized in Chapter IV of this document. Relevant information from
that chapter is also presented below.
Draft -Do not cite or quote C - 2 February 20, 2002
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Occurrence Data for Water
Adequate data are available to estimate central tendency and high-end values for exposure
to DBCM from treated surface and ground water. Numerous studies (summarized in Chapter IV
of this document) have examined the levels of DBCM in disinfected water. Of these studies, the
Information Collection Rule (ICR) data (U.S. EPA, 2001) most closely met the requirements for
sample size, geographic representation, reporting of analytical limits, and relevance to current
conditions. This survey examined the occurrence of brominated trihalomethanes in public water
supplies (PWSs) serving at least 100,000 persons as required by the Information Collection Rule
promulgated by U.S. EPA in May of 1996 for disinfectants and disinfection byproducts
(D/DBPs). The rule covered both surface and ground water systems. Monitoring data were
collected from about 300 water systems operating 501 plants over thel8-month period between
July 1997 and December 1998. At each plant, samples were collected monthly and analyzed for a
variety of D/DBPs on a monthly or quarterly basis. DBCM was among the analytes evaluated
quarterly (U.S. EPA, 2001). Five samples were taken each quarter at each plant - one of the
finished water and four of the water in the distribution system. Of the four samples from the
distribution system, one represented a sample with the same residence time as a finished water
sample held for a specific period of time, two represented approximate average water residence
times in the system, and one sample was taken where water residence time in the system is the
longest. For each plant and reporting period, EPA compiled several summary statistics. The
Distribution System (DS) Average value is the average of the four distribution system samples.
The DS High Value is the highest concentration of the four distribution system samples collected
by a plant in a given quarter. The DS High Value might be from any of the four samples and
could vary from quarter to quarter depending on which sample yielded the highest concentrations
in each quarter (U.S. EPA, 2001a). Table C-l summarizes the results of all six of the quarterly
reporting periods. The DS average and 90th percentile values for DBCM in surface water were
4.72 |ig/L and 5.57 |ig/L, respectively. The DS average and 90th percentile values for
dibromochloromethane in groundwater water were 3.09 |ig/L and 8.94 |ig/L, respectively.
U.S. EPA set a minimum reporting level (MRL) for DBCM of 1.0 iigfL for the ICR. The
MRL is a level below which systems were not required to report their monitoring results, even if
there were detectable levels. Values below the MRL were assigned a value of zero for the
purpose of calculating averages; this assignment affects the calculation of mean values for finished
water and DS high results and calculation of all DS average values.
Data for Occurrence in Air
Occurrence data for DBCM in ambient outdoor air were available from three reports
(Brodzinsky and Singh, 1983; Shikiya et al., 1984; Atlas and Shauffler, 1991). Brodzinsky and
Singh (1983) reviewed and summarized existing data for DBCM concentrations in ambient
outdoor air for several urban/suburban or source dominated locations across the United States
(Table C-2). No concentration data were available for rural or remote areas. The authors
reported mean, median, first and third quartile values, and minimum and maximum values by city.
In addition, they reported the same measures when the data were grouped by type of location
(i.e.,
Draft -Do not cite or quote C - 3 February 20, 2002
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Table C-l DBCM Concentrations Measured in U.S. Public Drinking Water Systems
Serving 100,000 or More Persons
Source
Data Type"
Number of
Samples
Medianb
Meanb
90th
Percentile
Range
DBCM (Mg/L)
Surface
Water
Finished
1853
1.9
4.03
12.0
<1.0 - 55.1
DS Average
1655
2.40
4.72
13.2
0 - 67.3
DS High
1655
2.9
5.57
15.0
< 1.0 -
67.3
Ground
Water
Finished
604
< 1.0
1.38
4.10
<1.0 - 33
DS Average
602
1.35
3.09
8.94
0 - 37.5
DS High
602
2.1
4.60
12.9
<1.0 - 85
Source: Disinfectants and Disinfection Byproducts (D/DBPs) ICR Data, U.S. EPA (2001).
a Finished = sample location after treatment, before entering the distribution system (DS); DS Average =
average of four sample locations in the DS; DS High = the highest concentration of the four
distribution system samples collected by a plant in a given quarter. For purposes of calculations,
all values below the minimum reporting level (MRL) of 1.0 , 'g/L for all three compounds were assigned a
value of zero.
b Median and mean of all samples including those below the MRL.
urban/suburban or source dominated), and when all data were combined. Dibromochloromethane
was detected in the air samples from Magnolia, AR, El Dorado, TX, Chapel Hill, NC, Beaumont
TX, and Lake Charles, LA at mean concentrations of 0 ppt, 0.48 ppt, 14 ppt, 14 ppt, and 19 ppt,
respectively. Data from these sites were combined for additional statistical analyses. The study
authors indicated that a value of 0.0 was entered for samples below the detection limit. The
detection limits from individual studies were not reported. Mean (± standard deviation) outdoor
air concentrations in urban/suburban and source dominated locations, respectively, were 15 ± 4
ppt and 0.28 ± 0.67 ppt for DBCM. Brodzinsky and Singh (1983) also calculated overall (grand)
means based on data from all sites. The grand mean value for DBCM was 3.8 ppt (n = 89, with
63 nondetects). When expressed on a ng/m3 basis, the corresponding mean value was 0.032
|ig/m3. Assuming an inhalation rate of 20 m3/day, this concentration results in a daily intake of
0.6 |ig/day. Assuming a rate of 13.2 m3/day, this concentration results in a daily intake of 0.43
|ig/day.
Shikiya et al. (1984) analyzed ambient air samples collected at four urban/industrial
locations in the California South Coast Air Basin from November 1982 to December 1983 for the
presence of DBCM. The sampling locations were El Monte, downtown Los Angeles,
Dominguez, and Riverside. The air samples were analyzed using gas chromatography with
detection by electron capture. The quantitation limit, defined as a level 10 times greater than the
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noise level, was 10 ppt by volume. The detection limit was defined as three times the noise level.
Most data in
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Table C-2 Selected Concentration Data for Individual Brominated Trihalomethanes (ppt)
in Outdoor Air as Summarized in Brodzinsky and Singh (1983)a'b
City
n
Non-
detects
Mean
(Std dev.)
Median
3rd
Quartil
e
Maximum
Reference
DBCM
Individual Sites
Beaumont, TX
11
0
14 (0.0)
14
14
14
Wallace (1981)
Chapel Hill, NC
6
0
14 (0.0)
14
14
14
Wallace (1981)
El Dorado, AR
40
35
0.48
(0.82)
0.0
0.82
2.5
Pellizzari et al.
(1978)
Lake Charles, LA
4
0
19 (9.6)
21
27
27
Pellizzari (1979)
Magnolia, AR
28
28
0.0 (0.0)
0.0
0.0
0.0
Pellizzari et al.
(1978)
Totals
Urban/Suburban
21
0
15 (4.2)
14
14
27
-
Source Areas
68
63
0.28
(0.67)
0.0
0.0
2.5
-
Grand Totals
89
63
3.8 (6.7)
0.0
2.5
27
-
" Includes only data considered to be of adequate, good, or excellent quality by the study authors.
b Concentrations are reported as parts per trillion by volume
this report were presented graphically. A few additional details were presented in a short
summary statement for each chemical. Summary data for each compound included monthly
means and composite means. The monthly means were calculated as the average of all data at a
site that were above the quantitation limit for a single month; samples with concentrations below
the limit of detection were not included in the calculations. The composite means were calculated
as the average value of all data for each compound above the quantitation limit at each site. Only
seventeen percent of the samples had DBCM levels above the quantitation limit of 10 ppt (0.085
|ig/m3). The highest reported concentration, monthly mean, and mean composite for DBCM were
290 ppt (2.5 |ig/m3), 280 ppt (2.4 |ig/m3), and 50 ppt (0.43 |ig/m3), respectively; all were
recorded in downtown Los Angeles in June. Only two monthly means were above 160 ppt; the
remainder of the monthly means were below 60 ppt.
Atlas and Schauffler (1991) collected replicate air samples at various locations on the
Island of Hawaii during a month-long field experiment to test an analytical method for
determining halocarbons in ambient air. DBCM was found at a mean level of 0.27 ppt. This
information was obtained from a secondary source which did not report the detection limit.
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Wallace et al. (1982) conducted a pilot study designed to field test personal air-quality
monitoring methods. Personal air samples were collected from students at two universities:
Lamar University, Texas, located near a petrochemical manufacturing area, and the University of
North Carolina (UNC), located in a nonindustrialized area. The samples were analyzed for a
number of volatile organic compounds, including brominated trihalomethanes. DBCM was not
detected at either location. Based on an analytical limit of 0.12 |ig/m3 or 0.018 ppb, these data
suggest that exposure via personal air is less than 2.4 |ig/day.
There are several limitations associated with the available data on occurrence of DBCM in
outdoor air. The available studies were collectively limited to five states (Arkansas, California,
Hawaii, North Carolina, and Texas). With the possible exception of the Hawaiian study (Atlas
and Schauffler 1991), all data were collected from urban/suburban or source dominated locations.
Thus, the data from these studies are not considered to be geographically representative of the
United States. In addition, sample size was not explicitly reported in the Shikiya et al. (1984)
study and the reported means were based only on data above the detection limit (only 17% of
total samples). An independent statistical evaluation could not be performed because raw data
were not presented. The data presented by Brodzinsky and Singh (1983) were obtained from
multiple sources and combined results for sampling periods ranging from instantaneous grab
samples to 24 hour averages (Wallace, 1997). The data from the Shikiya et al. (1984) and
Brodzinsky and Singh (1983) reports are approximately 20 years old and may not accurately
reflect current conditions.
Relatively few studies have reported the concentrations of trihalomethanes in indoor air of
homes. Kostiainen (1995) identified over 200 volatile organic compounds in indoor air of 26
houses identified by residents as causing symptoms such as headache, nausea, irritation of the
eyes, drowsiness, and fatigue. DBCM was not reported among the detected compounds.
Weisel et al. (1999) measured brominated trihalomethane concentrations in indoor air in
49 New Jersey residences selected to represent low and high levels of drinking water
contamination with trihalomethanes. Descriptive statistics for DBCM concentration in water
were provided for the combined high and low concentration groups, but not for the individual
groups. One valid 15-minute air sample was collected at each of 48 residences. The indoor air
concentrations of DBCM averaged 0.44 ± 0.95 |ig/m3 (0.052 ±0.11 ppb) and 0.53 ± 0.84 |ig/m3
(0.062 ±0.09 ppb) in the low and high water concentration residences with detection frequencies
of 5/25 and 7/23, respectively. The detection limit was 0.14 |ig/m3 (C. Weisel, personal
communication). It was not clear whether the averages were based on all measured samples or
only those samples that were above the detection limit. For this reason, the data were not used
for calculation of exposure to DBCM from indoor air.
It is possible that DBCM has been surveyed in studies of volatile organic compounds in air
and not reported because it was below detection limits. This has been suggested by Dr. Joachim
Pleil of the U.S. EPA Office of Research and Development, who is highly experienced in air
monitoring of volatiles including trihalomethanes. According to Dr. Pleil, the analytical methods
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used in analysis of volatile organic compounds are sufficiently sensitive to detect DBCM even if is
present only in minute quantities. For example, a survey of volatile organic compounds in indoor
air by Pleil et al. (1985) using EPA Method TO-14 (detection limit approximately 100 ppt by
volume, or 0.85 |ig/m3, for trihalomethanes) would certainly have detected DBCM had it been
present (J. Pleil, personal communication).
To accurately estimate total daily inhalation exposures from indoor and outdoor air, the
following data needs to be evaluated: location and season, the time spent indoors compared with
outdoors, potential exposures of individuals while showering or bathing, potential exposure from
volatilization of DBCM during other household activities (e.g., use of dishwashers, toilet
flushing), exposures of individuals who spend large amounts of time at indoor pools or in hot
tubs, and potential for occupational exposures (e.g., for laundromat or sewage treatment plant
workers). The existing measurement data are not adequate for such a refined analysis, but may be
used to roughly estimate intake from outdoor air.
Data for Occurrence in Food
Information on the levels of DBCM in foods and beverages is limited. Chlorine is used in
food production for applications such as the disinfection of chicken in poultry plants and the
superchlorination of water at soda and beer bottling plants (Borum, 1991). Therefore, the
possibility exists for contamination of food from chlorination by-products in foods with resulting
dietary exposure.
Two studies have reported analyses of commercial beverages for DBCM. In Italy,
Cocchioni et al. (1996) analyzed 61 samples of different commercially prepared beverages and 94
samples of mineral waters for volatile organo-halogenated compounds. Maximum DBCM
concentrations of 13.9 |ig/L (ppb) were found in prepared beverages, with a frequency of
detection of 43% (26/61), with a detection limit of less than 1 |ig/L (ppb). McNeal et al. (1995)
examined 27 different prepared beverages and mineral waters in the United States for DBCM at a
detection limit of 0.1 ng/g (ppb). DBCM was detected at 1 ng/g (ppb) in only one of seven types
of mineral and sparkling waters examined. DBCM was not detected in any of 5 flavored
noncarbonated beverages examined. DBCM was detected in only 4 of the 13 carbonated soft
drinks examined at levels of 0.5 to 2 ng/g (ppb). DBCM was not detected in either of the two
types of beer examined.
Two studies have tested for DBCM in individual food items. McNeal et al. (1995) tested
several types of food products and water from canned vegetables in the United States for DBCM.
DBCM was not detected in any of the samples. The foods examined included two types of
canned tomato sauce, canned pizza sauce, canned vegetable juice, vegetable waters from two
types of canned green beans and one type of sweet corn, duck sauces, beef extract, and Lite syrup
product. Imaeda et al. (1994) examined bean curd commercially available in Japan for
trihalomethanes. DBCM was not found in any of ten samples analyzed at a detection limit of 0.1
ppb.
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Kroneld and Reunanen (1990) analyzed pasteurized and unpasteurized cow's milk for
DBCM content in a study conducted in Turku, Finland. DBCM was detected in only one sample
of pasteurized milk at 5 |ig/L (ppb). The detection limit was not specified and information sample
size was unavailable in the secondary source that reported this study (U.S. EPA, 1994). DBCM
was not detected in unpasteurized milk. The presence of the DBCM in pasteurized milk may have
resulted from the use of chlorinated water during processing.
Estimates for dietary intake of DBCM by residents of the United States were not identified
in the materials reviewed for this document. Information on the levels in U.S. foods is too limited
to independently calculate a reliable estimate. However, the available data suggest that the
concentration of DBCM in foods is low.
Data for dietary intake of DBCM are available from a study conducted in Japan. Toyoda
et al. (1990) analyzed the dietary intake of DBCM by 30 housewives in Nagoya and Yokohama.
Duplicate portions of daily meals were collected for three consecutive days, sampled for DBCM
and analyzed at a detection limit of 0.2 ppb. The amount and types of food consumed were not
reported. This omission prevents a comparison of the studied diet to that consumed by the U.S.
population. The concentration of DBCM in the Japanese diet ranged from undetectable to
0.6 ppb (average, 0.1 ±0.2 ppb), and the mean dietary intake was estimated to be 0.3 ± 0.3
|ig/day. These data are considered adequate only for a rough estimate of the dietary intake of
DBCM.
Evaluation
The occurrence data base for DBCM in tap water consists of nationally aggregated data
and is considered adequate for determination of the RSC. In comparison, fewer occurrence data
are available for DBCM in food and outdoor air. The available air and food occurrence data,
although limited, permit rough estimates of intake.
Determination of the RSC
The RSC is calculated as follows:
RSC =
DI„
DI,
total
(1)
where:
DL
DI
total
^ I water, ingestion water, inhalation water, dermal
water, total "^^^outdoor air ^^food
(2)
(3)
The estimation of individual terms in these equations is described below.
Exposure Associated with Tap Water Uses
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Exposure to DBCM as a chlorination by-product in residential water can occur via three
primary exposure routes: 1) by ingestion; 2) by inhalation of DBCM volatilized during use of tap
water for bathing, showering, and other household activities; and 3) by dermal exposure during
showering and bathing. The existence of these routes for DBCM is supported by recent study
data. Kerger et al. (2000) demonstrated that levels of DBCM in indoor air are related to the use
of tap water for showering and bathing. Increases in the level of DBCM in the breath or blood
after showering or bathing have been documented in human subjects (Weisel et al., 1999; Backer
et al., 2000; Lynberg et al., 2001). Quantitative estimates of average daily exposure from
volatilized DBCM or dermal contact have been calculated and described in the following pages
for comparison with other routes of intake. These derived values have solid scientific support and
are sufficient for a reliable estimate. It is important to note that these estimates are for exposure
of the general population via tap water. Individuals who participate in activities such as
swimming or hot tub use may experience increased dermal and or inhalation uptake or brominated
trihalomethanes as a result of increased contact time with disinfected water. It is important to
note that water in hot tubs and swimming pools is routinely subjected to additional disinfection
and may not be representative of tap water using for drinking, cooking, and other household
activities.
a. Ingestion of DBCM in Drinking Water
Ingestion of DBCM is calculated by multiplying an appropriate intake rate for tap water
by the concentration of the compound found in tap water. Mean water intake rates of 1.2 and 0.6
L/day (NRC, 1999) were used for total mean ingestion from all uses and for direct ingestion (i.e.,
direct ingestion from tap, does not include use of tap water for making coffee and tea, soup, etc.),
respectively. An adjustment for intake of commercial beverages (e.g., soft drinks or mineral
waters) was not applied, because the intake of these beverages would be subtracted from the daily
tap water intake and the available data (e.g., McNeal, 1995) suggest that the level of DBCM in
such beverages is usually less than or similar to the level found in tap water. Assuming a tap
water concentration of 4.72 |ig/L (the distribution system average for DBCM in treated surface
water), the total and direct intakes of DBCM via ingestion of tap water are 5.7 |ig and 2.8 |ig,
respectively.
b. Inhalation of Waterborne DBCM
A three-compartment model approach was used to investigate the exposure from water-
related dibromochloromethane in indoor air. The three-compartment model employed was that of
McKone (1987). This model predicts the concentration of a volatile chemical in water (in this
case, DBCM) in each of three compartments of a house: the shower, the bathroom, and the
remainder of the house. The three-compartment model recognizes that most household water uses
are episodic rather than continuous, and room barriers (walls, doors) may restrict the rapid mixing
of DBCM released into air in one location with whole-house air, leading to occasional high levels
of DBCM in some rooms (especially those with high water usage, such as the shower or laundry).
Because concentrations are not constant, results are calculated as a function of time throughout
the day. Based on the time- and compartment-specific concentration values, human exposure
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levels in each compartment can then be calculated based on an assumed pattern of human
occupancy and behavior within the house. McKone (1987) estimated the source term for the
release of VOCs from water to air in each of the three compartments by extrapolation from
measurements of radon release using the VOC-specific Henry's law constant and the liquid- and
gas-phase diffusion coefficients. Basic equations and inputs to the model are provided below:
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Transient three-compartment model based on transfer efficiency developed
by McKone (1987)
V.
dyf (t)
dt
s,(»)
WUtfS, )-TE';oc
*¦* *-°
T — T
i i
rj-tJjVOC JTg*/CM _ " "~(JL ' "VUC rj-tj^Kn
i ' HT A ' (
OL Rn
K A
T Rn ^OL VOC
2.5 1
+
», yDl" H D'n
'g J
Rn
2.5 1
+
\
kD2l13 H'D™j
v L G y voc
Emh=^y,(t)OF,(t)dt
D°Se = E,nh ¦ BR
Where:
A interfacial area existing between water and air (cm2)
BR breathing rate (L/min)
C aqueous-phase concentration (mg/L)
Dl liquid-phase diffusion coefficient (cm2/sec)
Dg gas-phase diffusion coefficient (cm2/sec)
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ED
exposure duration (min)
Fink
inhalation exposure (mg/L, except for radon, which is in pCi/L)
F(t, T,°, T*)
function = 1 when time t is between x"and x*, and 0 otherwise
H
Henry's law constant (mg/La/mg/L,^)
Kol
overall mass-transfer coefficient (cm/min)
kola
overall interfacial mass-transfer coefficient (L/min)
OF it)
occupancy factor for compartment i at time t (1 if present, 0 if absent)
Q*
ventilation rate from compartment i to compartment j (L/min)
S,
emission rate from source in compartment i (mg/min)
TE{
transfer efficiency of chemical j during water use in compartment i (1)
r,0
time when water device in compartment i starts (min)
*
T,
time when water device in compartment i ends (min)
v,
volume of compartment i (L)
WU/rf, t/J
volume of water used in compartment i between time x/1 and x;* (L)
y,
gas-phase concentration in compartment i (mg/L)
Rn
radon
voc
volatile organic compound
The following properties of dibromochloromethane at 20°C were used as inputs to the
model:
H (mg/Lair/mg/Lwater) = 0.036
Dl (cm2/s) = 9.63 x 10"6
Dg (cm2/s) = 8.24 x 10"2
The following human activity and water use patterns were assumed. Four people living in
the house each take an 8-minute shower every morning, and spend 12 minutes in the bathroom
immediately thereafter. Each person spends an additional 20 minutes in the bathroom some time
during the remainder of the day. The second person taking a shower is selected for the purpose of
comparing exposure estimates. The person being modeled is assumed to spend 75% their time
indoors by assigning an average daily occupancy factor of 0.75.
The following specific parameters were used to implement the calculations of the three-
compartment model developed by McKone (1987):
Variable
Description
Value
PNUM
Number of people in the house
4
Person
Designated person for exposure calculation
2nd showerer
vs
Volume of shower
2000 L
Vb
Volume of bathroom
10000L
Va
Volume of main house
400000 L
Is
Volume of water used in shower
350 L
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Variable Description
Value
lb
Volume of water used in bathroom
350 L
la
Volume of water used in main house
450 L
Rs
Shower air residence time
3 min
Rb
Bathroom air residence time
12 min
Ra
Main house air residence time
78 min
Qsb
Ventilation rate from shower to bathroom
40000 L/hr
Qbs
Ventilation rate from bathroom to shower
40000 L/hr
Qab
Ventilation rate from main house to bathroom
50000 L/hr
Qba
Ventilation rate from bathroom to main house
45000 L/hr
Qbo
Ventilation rate from bathroom to outside
5000 L/hr
Qao
Ventilation rate from main house to outside
258000 L/hr
SFR
Shower flow rate
11 L/min
TER"
TE from shower to air for radon
0.7
TE^
TE from bathroom to air for radon
0.3
TE/"
TE from household water to air for radon
0.66
LS
Time in shower
8 min
h
Time in bathroom after shower
12 min
j i
lb
t 0
*
Ls
+ 0
Lb
*
h
t0
*
La
OF
Time in bathroom during rest of day
Time when first shower water use starts
Time when first shower water use ends
Time when toilet water use starts
Time when toilet water use ends
Time when other household water use starts
Time when other household water use ends
Daily average occupancy factor
20 min
7:00 a.m.
7:08 a.m.
12:00 a.m.
12:00 a.m.
7:00 a.m.
11:00 p.m.
0.75
BR
Breathing rate
9.2 L/min
Source for BR is U.S. EPA (1995) and for all other values is U.S. EPA (1993).
Based on these assumptions and parameter values, the model predicts an inhaled dose for DBCM
of 540 |ig/year per |ig/Lwatcr., Dividing this number by 365 days/year and multiplying the result by
4.72 |ig/L (the distribution system average for DBCM in treated surface water) gives an estimate
of 7 |ig/day for inhalation intake of DBCM volatilized from tap water.
c. Dermal Absorption of DBCM via Bathing or Showering
Estimates of dermal uptake of DBCM were obtained using a membrane model approach
as described in Cleek and Bunge (1993) and Bunge and McDougal (1998). The approach
assumes that the skin is composed of stratum corneum and viable epidermis. If the concentration
of the vehicle remains constant and the systemic concentration remains small, the dermal
absorption of chemicals can be divided into two periods: non-steady state and steady state. In the
non-steady state period, the chemical is absorbed in the lipophilic stratum corneum. The viable
dermis acts like a sink for the chemical once the steady-state is achieved. Equations are available
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for calculation of uptake under both steady state and non-steady state conditions. Because the
duration of exposure to DBCM during showering and bathing (10 minutes; U.S. EPA, 1992) is
substantially less than the time to steady state for DBCM absorption through skin (3.9 hours; U.S.
EPA; 1992), the non-steady state approach is the default procedure for estimating uptake:
Ik k , l t
D1- = 2ACJ " " """ (4)
where:
DA
event
= the amount of DBCM (|ig) absorbed in a 10-minute shower or bath
A
= the surface area exposed during the bath or shower event = 20,000 cm2 (U.S. EPA,
1992)
c„
= concentration of DBCM in tap water = 4.72 x 10"3 |ig/cm3 (U.S. EPA, 2001)
KP
= permeability coefficient = 3.9 x 10"3 cm/hr (U.S. EPA, 1992)
sc/w
= stratum corneum-water partition coefficient = 38 (see equation 5)
Lsc
= diffusion length of the stratum corneum = 1.5 x 10"3 cm (Bunge, personal
communication)
^event
= duration of exposure event = 10 minutes = 0.1667 hours (U.S. EPA, 1992)
The stratum corneum-water partition coefficient, Ksc/w, was estimated as recommended by
Bunge and McDougal (1998):
K = K°mn = 38 (5)
where
Kint = octanol water partition coefficient = 170 for DBCM (U.S. EPA, 1992)
When calculated using equation (4) and the input values above, the dermal uptake of DBCM
during a 10-minute showering or bathing event is approximately 0.65 |ig.
Perhaps the most difficult part of estimating dermal exposure is the determination of the
permeability coefficient, Kp, in the equation above. Estimates have been obtained from skin
penetration experiments. Most skin penetration experiments fall into one of two categories: in
vivo experiments performed on living humans or animals, and in vitro experiments made in
diffusion cells with excised skin from humans and animals. Determination of permeability
coefficients in vitro and in vivo generally requires that the exposure concentrations and surface
area are known and consistent. There is scientific debate over whether in vitro or in vivo
measurements are the most appropriate way to measure absorption of chemicals. In vitro
methods can provide quick and direct measures of flux and permeability coefficients. It is also
advantageous that human skin can be used in vitro when chemicals would be too toxic in in vivo
studies. Although the actual in vitro experiments are simpler, their use includes many important
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variables (e.g. animal species, thickness of skin, use of fresh or frozen skin, and receptor solution)
and uncertainties which influence the representativeness of the data. In vivo studies are often
more elaborate and require more data analysis.
The U.S. EPA (1992) Dermal Exposure Assessment document used in vitro data to
estimate the permeability coefficient (Kp) for multiple chemicals. An equation developed by Potts
and Guy (1993) was used to estimate the Kp of over 200 chemicals, including DBCM, as a
function of the octanol: water partition coefficient (K0J and molecular weight (MW). The
equation was derived from an experimental data base compiled by Gordon Flynn (1990), which
includes data for in vitro dermal absorption of about 90 chemicals from water:
log KP = -2.72+ 0.71(logA™)- 0.006MW (6)
where,
log Kow = logarithm of the octanol water partition coefficient = 2.23 for DBCM (U.S. EPA,
1992)
MW = molecular weight = 208.28 for DBCM
Despite the adequate correlations for representing experimental permeability data for a
broad rage of chemicals, experimental data may deviate from predictions made using the Potts and
Guy equation by one to two orders of magnitude. This variability was clearly demonstrated by
Vecchia (1997). Although uncertainty in experimental temperature and other data are partly
responsible, other known/unknown factors may also contribute to this discrepancy. For example,
the correlation assumes that MW is a good predictor for molecular size. This assumption may not
be appropriate for groups of compounds with chemical diversities affecting molecular size.
Halogenated hydrocarbons will occupy the same molar volume as a hydrocarbon molecule with a
much lower MW. As a result, equations based on MW that are developed from databases
consisting primarily of hydrocarbons will tend to systematically underestimate permeability
coefficients for chemically dense compounds such as DBCM, by an order of magnitude or perhaps
even more for compounds with specific gravity values larger than about 2.5 and MW greater than
200.
The Kp was calculated using the procedure of Vecchia and Bunge (2002) to address
potential underestimation of the Kp for DBCM by the prediction method used in U.S. EPA
(1992). This calculation used a modification (equation 7) of the Potts and Guy equation
(equation 6) which incorporates an adjustment factor for density of halogenated compounds when
compared to non-halogenated hydrocarbons:
MWOkc ,n,
log KP = -2.72 + 0.71(log IL) - 0.006 — (7)
j^hdo
where:
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logK^ = logarithm of the octanol water partition coefficient = 2.23 for DBCM (U.S. EPA, 1992)
MW = molecular weight = 208.28 for DBCM
phc = estimated liquid density of the compounds in the Flynn equation upon which equation
(6) was developed = 0.9 (Vecchia and Bunge, 2002)
Phaio = density of a halogenated hydrocarbon = 2.38 for DBCM (U.S. EPA, 1994)
The resulting value for Kp is 0.03 cm/hr. Substitution of this adjusted value in equation (4)
results in an uptake estimate of 2.0 |ig for a 10-minute showering or bathing event This value is
about 3-fold higher than the estimate obtained for dermal absorption using the unadjusted Kp
value reported in U.S. EPA (1992). The 2 |ig value was selected for determination of the RSC
for DCBM.
Exposure Associated Via Food and Outdoor Air
Dietary intake data for DBCM from a Japanese study were used in the absence of intake
data for U.S. residents. These data were used with the understanding that the composition (and
thus levels of DBCM) of U.S. and Japanese diets may differ. The grand mean for outdoor air
concentration calculated by Brodzinsky and Singh (1983) was used to estimate intake from
outdoor air. Although these data for food intake and air concentrations have limitations, they
were considered sufficient for a screening level estimate of the RSC.
Calculation of the RSC
An example calculation of the RSC is shown below. Results of calculations using different
exposure assumptions are summarized in Table C-3.
^ I water ^ I water, ingestion water. inhalation water. dermal
= 5.7 |ig/day + 7.0 ug/day + 2.0 |ig/day =14.7 |ig/day
DItotal = DIair + DIfood + DIwater (3)
= 0.4 |ig/day + 0.3 |ig/day + 14.7 |ig/day
= 15.4 |ig/day
RSC
DI,
DI,
total
14.7 |ig
15.4ng
0.96 x 100
= 95%
where:
DIwater = Intake of DBCM from tap water
DItotai = Intake of DBCM from all relevant sources (i.e., water, air, and food for DBCM)
DIair = Intake of DBCM from outdoor air, assuming 0.032 |ig/m3 and intake of 13.2 m3/day
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Dlfood = Intake of DBCM from food = 0.3 |ig/day after Toyoda (1990)
Dlwater, ingestion = Intake of DBCM from tap water by ingestion (assumes intake of 1.2 L/day)
DIWater> inhalation = Intake of DBCM from tap water by inhalation, as determined using 3-
compartment model; assumes intake of 13.2 m3/day (U.S. EPA, 1995)
D^ater, dermal = Dermal absorption of DBCM during bathing or showering using adjusted value of
Kj,; assumes 1 shower or bath/day
The RSC calculated using the ICR distribution system mean for surface water and U.S.
EPA default values for inhalation and drinking water ingestion was 0.95 or 95%. Substitution of
groundwater concentration data and/or the direct value for drinking water intake also resulted in
RSC values greater than 90%. These calculations suggest that an RSC as high as the default
ceiling of 80% is justified for DBCM.
The uncertainty in use of 80% for the RSC is related to the quality of data for intake from
food and outdoor air. The primary concern is that the available data might result in a significant
underestimate of the actual exposure via these media, resulting in an RSC value that was not
appropriately protective of health. A series of calculations was performed to test the effect of
underestimating exposure from outdoor air and/or food on the RSC. An arbitrary 10-fold
increase in the dietary intake of DBCM while holding intake from other sources constant resulted
in RSC values of 78% or 81%, depending upon the intake values selected. Increasing the intake
of DBCM from outdoor air by 10-fold resulted in RSC values of 72% or 76%. Increasing both
food and air intake of DBCM by 10-fold gave RSC values of 62% or 67%. These calculations
assumed a tap water concentration of 4.72 |ig/L, which is the Information Collection Rule
distribution system mean for surface water (U.S. EPA, 2001). These calculations suggest that an
RSC of 80%) would be protective of human health even if concentrations of DBCM in food or
outdoor air were underestimated by a factor of almost 10-fold.
Conclusion
An RSC of 80%) is recommended based on the analysis of available data on occurrence of
DBCM in tap water and other media. The major uncertainties in this analysis are related to
limited measurement data for DBCM in outdoor air and in food.
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C- 18
February 20, 2002
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Table C-3 Results of RSC Calculations for DBCM
Water Source
cw
(Hg/L )a
Condition
TP
lxvwater
(L/day)b
IRair
(m3/day)c
Dlfood
(|ig/day)d
DLur
(Hg/day)e
water
(Hg/day)f
DItotal
(Hg/day)g
RSC
(%)
Surface
4.72
Surface Water
0.6
13.2
0.3
0.42
11.8
12.6
94
1.2
13.2
0.3
0.42
14.7
15.4
95
Ground
3.09
Ground Water
0.6
13.2
0.3
0.42
10.9
11.6
94
1.2
13.2
0.3
0.42
12.7
13.4
95
Surface
4.72
If intake of DBCM
from food
increased by 10-
fold
0.6
13.2
3.0
0.42
11.8
15.3
78
1.2
13.2
3.0
0.42
14.7
18.1
81
Surface
4.72
If intake of DBCM
from air increased
by 10-fold
0.6
13.2
0.3
4.2
11.8
16.4
72
1.2
13.2
0.3
4.2
14.7
19.2
76
Surface
4.72
If intake of DBCM
from air and food
increased by 10-
fold
0.6
13.2
3.0
4.2
11.8
19.1
62
1.2
13.2
3.0
4.2
14.7
21.9
67
a Concentration of DBCM in water: ICR distribution system mean (U.S. EPA, 2001).
b Daily intake rate for water. Values are for direct (0.6 L/day; NRC. 1999) or total mean (1.2 L/day; NRC, 1999) ingestion rates.
c Daily inhalation rate. Value used is consistent with the input value of 9.2 L/min for the three-compartment model used to estimate intake of DBCM from
indoor air.
d Daily intake of DBCM in food, based on data from Toyoda et al. (1990).
e Daily intake of DBCM in air, based on grand mean for DBCM concentration in outdoor air calculated by Brodzinsky and Singh (1983)
f Daily intake of DBCM in water from ingestion, inhalation of volatilized compound, and dermal absorption. See text for details of calculation.
B Total daily intake of DBCM from water, outdoor air, and food.
Draft -Do not cite or quote
C - 19 February 20, 2002
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References Cited
Atlas, E. and S. Schauffler. 1991. Analysis of alkyl nitrates and selected halocarbons in the
ambient atmosphere using a charcoal preconcentration technique. Environ. Sci. Technol.
25(l):61-67. (As cited in U.S. EPA, 1994.)
ATSDR. 1990. Agency for Toxic Substances and Disease Registry. Toxicological Profile for
Bromoform and Chlorodibromomethane. PB/91/180323/AS. Available online at
www. atsdr. cdc. gov/toxprofiles/tp 13 0. pdf
Backer, L.C., D.L. Ashley, M.A. Bonin, F. Cardinali, S.M. Kieszak, and J.V. Wooten. 2000.
Household exposures to drinking water disinfection by-products: whole blood trihalomethane
levels. J. Expo. Anal. Environ. Epidemiol. 10:321-326.
Borum, D. 1991. Telephone conversation (December 17) between Denis Borum, OST/OW,
U.S. Environmental Protection Agency, Washington, D.C., and Jeff Mosher, Wade Miller
Associates, Arlington, VA. (As cited in U.S. EPA, 1994.)
Brodzinsky, R. and H.B. Singh. 1983. Volatile organic chemicals in the atmosphere: An
assessment of available data. U.S. Environmental Protection Agency, Office of Research and
Development, Research Triangle Park, NC. (As cited in U.S. EPA, 1994.)
Bunge, A.L. and J.N. McDougal. 1999. Dermal Uptake. In: S.S. Olin (Ed.), Exposure to
Contaminants in Drinking Water: Estimating Uptake Through the Skin and By Inhalation.
Washington, D.C.: International life Sciences Institute.
Cleek, R.L. and A.L. Bunge. 1993. A new method for estimating dermal absorption from
chemical exposure. 1. General approach. Pharm. Res. 10:497-506., Bunge AL.
Cocchioni, M., M.G. Pellegrini, L. Nacciarriti, and G. Bernacchia. 1996. Organo-halogenated
compounds in commercial prepared beverages. II. Igiene Moderna. 106:287-293. [Italian],
Crump, K. 1995. Calculation of benchmark doses from continuous data. Risk Analysis 15:79-
89.
Flynn, G.L. 1990 .Physiochemical determinants of skin absorption. In: T.R. Gerity and C.J
Henry, eds. Principles of route-to-route extrapolation for risk assessment. Amsterdam: Elsevier
Science Publishing Co. Inc; pp. 93-127.
Imaeda, K., Yamamoto, D., Watanabe, T., Yoshimura, Y., Uchiyama, K., Kato, K., and Ohsawa,
K. 1994. Formation of trihalomethanes during production of bean-curd. Jap. J. Toxicol. Environ.
Health 40:527-533.
Draft -Do not cite or quote
C - 20
February 20, 2002
-------
Kerger, B.D., C.E. Schmidt, and D.J. Paustenbach. 2000. Assessment of airborne exposure to
trihalomethanes from tap water in residential showers and baths. Risk Anal. 20:637-651.
Kostiainen, R. 1995. Volatile organic compounds in the indoor air of normal and sick houses.
Atmos..Environ. 29:693-702.
Kroneld, R. and M. Reunanen. 1990. Determination of volatile pollutants in human and animal
milk by GC-MS. Bull. Environ. Contam. Toxicol. 44(6):917-923. (As cited in U.S. EPA, 1994.)
Lynberg, M., J.R. Nuckols, P. Langois, D. Ashley, P. Singer, P. Mendola, C. Wilkes, H. Krapfl,
E. Miles, V. Speight, B. Lin, L. Small, A. Miles, M. Bonin, P. Zeitz, A. Tadkod, J. Henry, M.B.
Forrester. 2001. Assessing exposure to disinfection by-products in women of reproductive age
living in Corpus Cristi, Texas, and Cobb County, Georgia: descriptive results and methods. Env.
Health Perspect. 109(6):597-604.
McKone, T.E. 1987. Human Exposure to Volatile Organic Compounds in Household Tap
Water: The Indoor Inhalation Pathway. Environ. Sci. Technol. 27(12), 1194-1201.
McNeal, T.P., H.C. Hollifield, and G.W. Diachenko. 1995. Survey of trihalomethanes and other
volatile chemical contaminants in processed foods by purge-and-trap capillary gas
chromatography with mass selective detection. J. AOAC Int. 78:391-397.
NRC. 1999. National Research Council. Risk Assessment of Radon in Drinking Water.
Washington, D.C.: National Academy Press.
Pellizzari, E.D. 1979. Organic screening in Lake Charles, LA, using gas chromatography mass
spectrometry computer techniques. EPA contract no. 68-02-2714. Research Triangle Institute,
Research Triangle Park, NC. (As cited in Brodzinsky and Singh, 1983).
Pellizzari, E.D., R.A. Zweidinger, and M.D. Erickson. 1978. Environmental monitoring near
industrial sites: Brominated chemicals Part I. Report Nos. NTIS PB286484 and EPA-560/6-78-
002. Prepared for the U.S. Environmental Protection Agency by the Research Triangle Institute,
Research Triangle Park, NC. (As cited in Brodzinsky and Singh, 1983).
Pleil, J., K. Oliver, and W.A. McClenny. 1985. Volatile organic compounds in indoor air: a
survey of various structures. In: Indoor Air Quality in Cold Climates: Hazards and abatement
measures. D. Walkinshaw, Ed. Transactions of the Air Pollution Control Association, p. 237-
247.
Potts R.O. and R.H. Guy. 1993. Predicting skin permeability. Pharm. Res.9:663-669.
Shacklette, H.T, and J.G. Boerngen. 1984. Element concentrations in soils and other surficial
materials of the conterminous United States. U.S. Geological Survey Professional Paper 1270.
Washington. D.C.: U.S. Government Printing Office.
Draft -Do not cite or quote
C - 21
February 20, 2002
-------
Shikiya, J., G. Tsou, J. Kowalski and F. Leh. 1984. Ambient monitoring of selected halogenated
hydrocarbons and benzene in the California south coast air basin. Proc-APCA 77th Annual Mtg.,
Vol. 1.
Toyoda, M., K. Takagi, A. Tsurumizu, Y. Saito, H. Kikawa, T. Kawamura and Y. Sakabe. 1990.
Estimation of daily intake of four kinds of trihalomethane by Japanese housewives. J. Food Saf.
10(3):191-200.
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drinking water. Report to
Congress, Washington, D.C.: U.S. EPA. EPA-56014-75-005. PB 260961. (As cited in U.S.
EPA, 1994.)
U.S. EPA. 1987. U.S. Environmental Protection Agency. Health effects assessment for
trihalogenated methanes. EPA/600/8-88/059.
U.S. EPA. 1992. United States Environmental Protection Agency. Dermal Exposure
Assessment: Principles and Applications. Interim Report. Exposure Assessment Group. Office of
Health and Environmental Assessment. EPA/600/8-91/01 IB. Washington, D.C. January.
U.S. EPA. 1993. United States Environmental Protection Agency. Uncertainty Analysis of Risk
Associated with Exposure to Radon in Drinking Water, EPA-TR-1656-3 A; U.S. Environmental
Protection Agency: Washington, DC.
U.S. EPA. 1994. U.S. Environmental Protection Agency. Final draft for the drinking water
criteria document on trihalomethanes. Contract no. EPA/68-C2-0139. Washington, D.C.: U.S.
Environmental Protection Agency Health and Ecological Division, Office of Science and
Technology, Office of Water.
U.S. EPA. 1995. United States Environmental Protection Agency. Uncertainty Analysis of Risks
Associated with Exposure to Radon in Drinking Water, EPA-822-R-96-005; U.S. Environmental
Protection Agency: Washington, D.C.
U.S. EPA. 1997. U.S. Environmental Protection Agency. Exposure Factors Handbook.
National Center for Environmental Assessment. PB98-124217. August.
U.S. EPA. 2001. U.S. Environmental Protection Agency. Stage 2 Occurrence and Exposure
Assessment for Disinfectants and Disinfection By-Products (D/DBPs) in Public Drinking Water
Systems.
Vecchia, B.E. 1997. Estimating the Dermally Absorbed Dose from Chemical Exposure: Data
Analysis, Parameter Estimation, and Sensitivity to Parameter Uncertainties, M.S. Thesis,
Colorado School of Mines, Golden, CO.
Draft -Do not cite or quote
C - 22
February 20, 2002
-------
Vecchia, B.E. and A.L. Bunge. 2002. Skin Absorption Databases and Predictive Equations.
Chapter 2, In: Transdermal Drug Delivery Systems, 2nd Edition. Eds. J. Hadgraft and R.H. Guy.
New York: Marcel Dekker.
Wallace, L. 1981. Measurements of volatile organic compounds in breathing-zone air, drinking
water, and exhaled breath (preliminary draft). (As cited in Brodzinsky and Singh, 1983).
Wallace, L., R. Zweidinger, M. Erickson, S. Cooper, D. Whitaker and E. Pellizzari. 1982.
Monitoring individual exposure: Measurements of volatile organic compounds in breathing-zone
air, drinking water, and exhaled breath. Environ. Intl. 8:269-282. (As cited in U.S. EPA, 1994.)
Wallace, L. A. 1997. Human exposure and body burden for chloroform and other
trihalomethanes. Crit. Rev. Environ. Sci. Technol. 27:113-194.
Weisel, C. P., H. Kim, P. Haltmeier, and J.B.Klotz. 1999. Exposure estimates to disinfection by-
products of chlorinated drinking water. Environ. Health Perspect. 107: 103-110.
Draft -Do not cite or quote
C - 23
February 20, 2002
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