United States
Environmental Protection
Agency
Health Effects Support
Document for
Hexachlorobutadiene
-------
Health Effects Support Document
for Hexachloro butadiene
U.S. Environmental Protection Agency
Office of Water (43 04T)
Health and Ecological Criteria Division
Washington, DC 20460
www.epa.gov/safewater/
EPA 822-R-03-002
February 2003
Printed on Recycled Paper
-------
FOREWORD
The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator of
the Environmental Protection Agency (EPA) to establish a list of contaminants to aid the agency in
regulatory priority setting for the drinking water program. In addition, SDWA requires EPA to make
regulatory determinations for no fewer than five contaminants by August 2001. The criteria used to
determine whether or not to regulate a chemical on the CCL are the following:
The contaminant may have an adverse effect on the health of persons.
The contaminant is known to occur or there is a substantial likelihood that the contaminant
will occur in public water systems with a frequency and at levels of public health concern.
In the sole judgment of the administrator, regulation of such contaminant presents a
meaningful opportunity for health risk reduction for persons served by public water systems.
The Agency's findings for the three criteria are used in making a determination to regulate
a contaminant. The Agency may determine that there is no need for regulation when a contaminant
fails to meet one of the criteria.
This document provides the health effects basis for the regulatory determination for
hexachlorobutadiene. In arriving at the regulatory determination, data on toxicokinetics, human
exposure, acute and chronic toxicity to animals and humans, epidemiology, and mechanisms of
toxicity were evaluated. In order to avoid duplication of effort, information from the following risk
assessments by the EPA and other government agencies was used in development of this document.
U.S. EPA. 199la. Drinking Water Health Advisory: Hexachlorobutadiene. In: Volatile
Organic Compounds. United States Environmental Protection Agency, Office of Drinking
Water. Lewis Publishers. Ann Arbor, Michigan.
ATSDR. 1994. Toxicological Profile for Hexachlorobutadiene. Agency for Toxic Substances
and Disease Registry, Department of Health and Human Services.
U.S. EPA, 1998a. Draft Ambient Water Quality Criteria for the Protection of Human Health.
Office of Water. EPA 822-R-98-004.
Information from the published risk assessments was supplemented with information from
recent studies of hexachlorobutadiene identified by literature searches conducted in 1999 and 2000
and the primary references for key studies.
Generally a Reference Dose (RfD) is provided as the assessment of long-term toxic effects
other than carcinogenicity. RfD determination assumes that thresholds exist for certain toxic effects
such as cellular necrosis. It is expressed in terms of milligrams per kilogram per day (mg/kg-day).
In general, the RfD is an estimate (with uncertainty spanning perhaps an order of magnitude) of a
daily exposure to the human population (including sensitive subgroups) that is likely to be without
an appreciable risk of deleterious effects during a lifetime.
HCBD February 2003 iii
-------
The carcinogenicity assessment for hexachlorobutadiene includes a formal hazard
identification as well as a quantitative dose-response assessment of the risk from oral exposure.
Hazard identification is a weight-of-evidence judgment of the likelihood that the agent is a human
carcinogen via the oral route and the conditions under which the carcinogenic effects may be
expressed.
Guidelines that were used in the development of this assessment may include the following:
the Guidelines for Carcinogen Risk Assessment (U. S. EPA, 1986a), Draft Guide lines for Carcinogen
Risk Assessment (USEP A, 1999c), Guidelines for the Health Risk Assessment of Chemical Mixtures
(U.S. EPA, 1986b), Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986c), Guidelines
for Developmental Toxicity Risk Assessment (U.S. EPA, 1991b), Proposed Guidelines for
Carcinogen Risk Assessment (1996a), Guidelines for Reproductive Toxicity Risk Assessment (U.S.
EPA, 1996b), and Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998b);
Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988); Use of the Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 1995);
Science Policy Council Handbook: Peer Review (U.S. EPA, 1998c); and Memorandum from EPA
Administrator, Carol Browner, dated March 21, 1995.
The chapter on occurrence and exposure to hexachlorobutadiene through potable water was
developed by the Office of Ground Water and Drinking Water. It is based primarily on unregulated
contaminant monitoring (UCM) data collected under SDWA. The UCM data are supplemented with
ambient water data as well as information on production, use, and discharge.
HCBD February 2003 iv
-------
ACKNOWLEDGMENTS
This document was prepared under the U.S. EPA Contract No. 68-C-02-009, Work Assignment
No. B-13 with ICF Consulting, Fairfax, VA. The Lead U.S. EPA Scientist is Diana Wong, Ph.D.,
DABT, Health and Ecological Criteria Division, Office of Science and Technology, Office of
Water.
HCBD February 2003
-------
TABLE OF CONTENTS
FOREWORD iii
ACKNOWLEDGMENTS v
LIST OF TABLES ix
LIST OF FIGURES x
1.0 EXECUTIVE SUMMARY 1-1
2.0 IDENTITY: PHYSICAL AND CHEMICAL PROPERTIES 2-1
3.0 USES AND ENVIRONMENTAL FATE 3-1
3.1 Uses 3-1
3.2 Release to the Environment 3-1
3.3 Fate in Air 3-3
3.4 Fate in Water 3-3
3.5 Fate in Soil 3-4
4.0 EXPOSURE FROM DRINKING WATER 4-1
4.1 Ambient Occurrence 4-1
4.1.1 Data Sources and Methods 4-1
4.1.2 Results 4-2
4.2 Drinking Water Occurrence 4-2
4.2.1 Data Sources, Data Quality, and Analytical Methods 4-3
4.2.2 Results 4-9
4.3 Conclusions 4-17
5.0 EXPOSURE FROM MEDIA OTHER THAN WATER 5-1
5.1 Exposure from Food 5-1
5.1.1 Concentrations in Non-Fish Food Items 5-1
5.1.2 Concentrations in Fish 5-1
5.1.3 Intake of HCBD from Food 5-3
5.2 Exposure from Air 5-4
5.2.1 Concentration of HCBD in Air 5-4
5.2.2 Intake of HCBD from Air 5-5
5.3 Exposure from Soil 5-6
5.3.1 Concentration of HCBD in Soil and Sediment 5-6
5.3.2 Intake of HCBD from Soil 5-6
5.4 Other Residential Exposures 5-6
5.5 Summary 5-6
6.0 TOXICOKINETICS 6-1
6.1 Absorption 6-1
HCBD February 2003 vi
-------
7.
7.
6.2 Distribution 6-2
6.3 Metabolism 6-2
6.4 Excretion 6-6
7.0 HAZARD IDENTIFICATION 7-1
7.1 Human Effects 7-1
7.1.1 Short-Term Studies 7-1
7.1.2 Long-Term and Epidemiological Studies 7-1
7.2 Animal Studies 7-2
7.2.1 Acute Toxicity 7-2
7.2.2 Short-Term Studies 7-6
7.2.3 Subchronic Studies 7-8
7.2.4 Neurotoxicity 7-10
7.2.5 Developmental/Reproductive Toxicity 7-11
12.6 Chronic Toxicity 7-13
12.1 Carcinogenicity 7-13
7.3 Other Key Data 7-15
7.3.1 Mutagenicity/Genotoxicity 7-15
7.3.2 Immunotoxicity 7-22
7.3.3 Hormonal Disruption 7-23
7.3.4 Physiological or Mechanistic Studies 7-23
7.3.5 Structure-Activity Relationship 7-27
7.4 Hazard Characterization 7-28
7.4.1 Synthesis and Evaluation of Major Noncancer Effects 7-28
7.4.2 Synthesis and Evaluation of Carcinogenic Effects 7-34
7.4.3 Mode of Action and Implications in Cancer Assessment 7-34
7.4.4 Weight of Evidence Evaluation for Carcinogenicity 7-36
7.4.5 Sensitive Populations 7-37
8.0 DOSE-RESPONSE ASSESSMENT 8-1
8.1 Dose-Response for Noncancer Effects 8-1
8.1.1 RfD Determination 8-1
8.1.2. RfC Determination 8-5
8.2 Dose-Response for Cancer Effects 8-5
8.2.1 Choice of Study 8-5
8.2.2 Dose-Response Characterization 8-6
8.2.3 Extrapolation Model and Rationale 8-10
8.2.4 Cancer Potency and Unit Risk 8-11
8.2.5 Discussion of Confidence 8-12
9.0 REGULATORY DETERMINATION AND CHARACTERIZATION OF RISK FROM
DRINKING WATER 9-1
9.1 Regulatory Determination for Chemicals on the CCL 9-1
9.1.1 Criteria for Regulatory Determination 9-1
9.1.2 National Drinking Water Advisory Council Recommendations 9-2
9.2 Health Effects 9-2
HCBD February 2003 vii
-------
9.2.1 Health Criterion Conclusion 9-3
9.2.2 Hazard Characterization and Mode of Action Implications 9-3
9.2.3 Dose-Response Characterization and Implications in Risk Assessment 9-4
9.3 Occurrence in Public Water Systems 9-8
9.3.1 Occurrence Criterion Conclusion 9-8
9.3.2 Monitoring Data 9-8
9.3.3 Use and Fate Data 9-9
9.4 Risk Reduction 9-10
9.4.1 Risk Reduction Criterion Conclusion 9-10
9.4.2 Exposed Population Estimates 9-11
9.4.3 Relative Source Contribution 9-11
9.4.4 Sensitive Populations 9-12
9.5 Regulatory Determination Summary 9-13
10.0 REFERENCES 10-1
APPENDIX A: Abbreviations and Acronyms A-l
APPENDIX B: Round 1 and Round 2 Occurrence Data Tables for Hexachlorobutadiene . . . B-l
HCBD February 2003 viii
-------
LIST OF TABLES
Table 2-1. Chemical and Physical Properties of Hexachlorobutadiene 2-2
Table 3-1. Environmental Releases (in pounds) for Hexachlorobutadiene in the United
States, 1988-1998 3-2
Table 4-1. Cross-section States for Round 1 (24 States) and Round 2 (20 States) 4-6
Table 4-2. Summary Occurrence Statistics for Hexachlorobutadiene 4-11
Table 5-1. HCBD Tissue Concentration in Fish Collected Near Four Chemical
Manufacturing Plants 5-3
Table 5-2. Summary of Concentration Data and Exposure Estimates for Media Other Than
Water 5-7
Table 7-1. Histopathological Findings in Adult Rats Fed Diets containing Hexachlorobutadiene.
7-9
Table 7-2. Mutagenicity of HCBD in Salmonella typhimurium Test Systems 7-17
Table 7-3. Mutagenicity of HCBD Metabolites 7-19
Table 7-4. Genotoxicity of HCBD in Eukaryotic Assay Systems 7-20
Table 7-5. Summary of Principal HCBD Toxicity Studies 7-29
Table 8-1. Incidence of Renal Tubular Regenerative Response in Mice Treated with HCBD for
13 Weeks 8-2
Table 8-2. Benchmark Dose Estimates from NTP (1991) Female Mouse Renal Tubular
Regeneration Response 8-3
Table 8-3. Incidence of Renal Tubular Neoplasms in Rats Treated with HCBD for 2
Years 8-6
Table 8-5. Summary of Cancer Risk Values for HCBD 8-12
Table 9-1. Dose-Response Information from Several Key Studies of HCBD Toxicity (Oral
Exposure) 9-5
Table 9-2. National Population Estimates for HCBD Exposure via Drinking Water 9-11
Table 9-3. Comparison of Average Daily Intakes from Drinking Water and Other Media a
9-12
Table 9-4. Ratios a of Exposures from Various Media to Exposures from Drinking Water.
9-12
HCBD February 2003 ix
-------
LIST OF FIGURES
Figure 2-1. Chemical Structure of Hexachlorobutadiene 2-1
Figure 4-1. Geographic Distribution of Cross-section States for Round 1 (left) and Round 2
(right) 4-6
Figure 4-2. States with PWSs with Detections of Hexachlorobutadiene for all States with
Data in URCIS (Round 1) and SDWIS/FED (Round 2) 4-14
Figure 4-3. States with PWSs with Detections of Hexachlorobutadiene (any PWSs with
results greater than the Minimum Reporting Level [MRL]) for Round 1 (above)
and Round 2 (below) Cross-section States 4-15
Figure 4-4. Cross-section States (Round 1 and Round 2 combined) with PWSs with
Detections of Hexachlorobutadiene (above) and concentrations greater than the
Health Reference Level (HRL; below) 4-16
Figure 6-1. Proposed Pathways for Hexachlorobutadiene Metabolism 6-4
Figure 8-1. Benchmark Dose Estimate Using Weibull Model 8-3
Figure 8-2. Renal Tumor Dose Response Curves 8-9
HCBD February 2003
-------
1.0 EXECUTIVE SUMMARY
The U.S. Environmental Protection Agency (EPA) has prepared this Health Effects Support
Document for Hexachlorobutadiene (HCBD) to assist in determining whether to regulate HCBD
with a National Primary Drinking Water Regulation (NPDWR). The available data on occurrence,
exposure, and other risk considerations suggest that, because HCBD does not occur in public water
systems at frequencies and levels of public health concern, regulating HCBD will not present a
meaningful opportunity for health risk reduction for persons served by public water systems. EPA
presents its determination and data analysis in the Federal Register Notice covering the
Contaminant Candidate List (CCL) regulatory determinations.
HCBD (Chemical Abstracts Services Registry Number 87-68-3) is a colorless liquid at room
temperature. It is poorly soluble in water, and has a high affinity for organic particulate matter.
HCBD has never been specifically manufactured as a commercial product in the United States.
However, significant quantities of hexachlorobutadiene are generated in the United States as waste
by-product from the chlorination of hydrocarbons. The chemical is used as an intermediate product
in rubber manufacturing and chlorofluorocarbon and lubricant production, as well as for transformer
and hydraulic fluids, fluid for gyroscopes, heat transfer liquid, solvents, laboratory reagents, and as
a wash liquor for removing C4 and higher hydrocarbons. Hexachlorobutadiene has also been used
as a fumigant in some overseas countries. Some of the chemical properties for hexachlorobutadiene
(CAS# 87-68-3) include the following: water solubility = 2-2.55 mg/L; vapor pressure (25°C) = 0.15
mmHg; Log Kow = 4.78; and Log Koc = 3.67.
Emissions into air is the major pathway of release. For hexachlorobutadiene, air emissions
constitute most of the on-site releases. Hexachlorobutadiene is listed as a toxic release inventory
(TRI) chemical. It is included in the Agency for Toxic Substances and Disease Registry's (ATSDR)
Hazardous Substance Release and Health Effects Database (HazDat) and has been detected in site
samples in fifteen States: AL, AZ, CA, CT, IA, LA, MI, MN, NJ, NY, OH, PA, RI, SC, WA
(ATSDR, 2000).
Ambient air concentration data are available from Shah and Heyerdahl (1988). The mean and
median of all ambient concentrations were 0.42 |ig/m3 and 0.04 |ig/m3, respectively. Air intake for
adults is estimated to be 1.2 x 10"4 mg/kg-day using the mean air concentration, and inhalation is the
main pathway of exposure (U. S. EPA, 1998a). Hexachlorobutadiene is not found in non-fish dietary
foods for the majority of regions in the US. It was detected in fish at 3% of 362 sites sampled. The
mean fish concentration at all sites was 0.6 ng/g (Kuehl et al., 1994). An estimate of adult exposure
via fish consumption is 1.54 x 10'7 mg/kg-day (U.S. EPA, 1998a).
Cross-sectional monitoring data from two rounds of sampling conducted under EPA's
Unregulated Contaminant Monitoring (UCM) program indicate that the frequency of detection of
HCBD in public water systems (PWSs) is low. Round 1, conducted from 1987 to 1992 in 24 States,
detected HCBD at levels above the minimum reporting level (MRL) of 0.5 |ig/L for 0.35% of the
PWSs, while Round 2, conducted from 1993 to 1997 in 20 States, detected HCBD at levels above
the MRL for 0.18% of the PWSs. The health reference level (HRL) of 0.9 |ig/L for HCBD is a
preliminary health effect level used in the UCM analysis (U.S. EPA, 2001c). It is the concentration
corresponding to 10"6 incremental cancer risk, calculated from the slope factor using the linear
HCBD February 2003 1-1
-------
method. In that analysis, 0.114% of the PWSs in Round 1 (74 systems) exceeded the HRL, while
0.018% (11 systems) in Round 2 exceeded the HRL. The United States Geological Survey's
National Ambient Water Quality Assessment (NAWQA) program did not detect HCBD in the
ground water or well water samples surveyed. When average daily drinking water intakes for HCBD
are compared with intakes from food, air and soil, drinking water accounts for a relatively small
proportion of total HCBD intake. This drinking water intake, however, does not include other uses
for potable water such as bathing and showering.
Hexachlorobutadiene (HCBD) is readily absorbed following oral administration in rats.
Following gastrointestinal uptake, HCBD and its metabolites distribute preferentially to the kidney,
liver, adipose tissue and the brain (Reichert et al., 1985). The primary pathway for HCBD
metabolism is conjugation with glutathione, with subsequent conversion to a cysteine conjugate.
Activation of the cysteine conjugate by P-lyase yields a highly reactive thioketene intermediate.
Covalent binding of this thioketene to DNA, proteins and other macromolecules is considered to be
the mechanism responsible for the observed cytotoxic and mutagenic effects of HCBD and its
metabolites. Evidence exists for a male specific metabolic pathway in rats (Birner et al., 1995). The
primary routes for elimination of absorbed HCBD are urinary and fecal excretion; a small amount
of absorbed HCBD is oxidized to carbon dioxide (U.S. EPA, 1991;U.S.EPA, 1999; ATSDR, 1994).
There are no reliable dose-response data for humans exposed to HCBD. There is no
information available to determine the carcinogenic potential of HCBD exposure in humans. Studies
in animals show a selective adverse effect of HCBD on the kidney, specifically the proximal tubule.
Subchronic (NTP, 1991) and chronic (Kociba et al., 1977) studies in rodents present a clear picture
of dose-related renal damage at 2 mg/kg-day and above, with possible effects at 0.2 mg/kg-day.
Progressive events over time include changes in kidney weight, renal tubular degeneration and
regeneration, hyperplasia, focal adenomatous proliferation, and renal tumor formation. One subacute
inhalation study also found enlarged adrenals and degeneration of adrenal cortex at 250 ppm (Gage,
1970). Developmental effects were also associated with hexachlorobutadiene exposure in animals
(Harleman and Seinen, 1979). However, these effects were observed at higher doses than required
to produce renal toxicity. Pups with lower birth weights and reduced body weight gain were reported
at maternal dose of 8.1-15 mg/kg-day in rats (Badaeva, 1983; Harleman and Seinen, 1979). In the
presence of metabolic activation, HCBD and its reactive metabolites are mutagenic in some
(Simmon, 1977; Reichert et al., 1984; Reichert and Schutz, 1986; Wild et al., 1986) but not all
studies. Only one study of lifetime oral exposure to hexachlorobutadiene was located (Kociba et
al., 1977). At 20 mg/kg-day, benign and malignant renal tumors were seen in approximately 23%
(9/39) of the male rats, and 15% (6/40) of the female rats. This dose exceeded the maximum
tolerated dose at which increased mortality, severe renal toxicity, and significant weight loss were
also observed. There were no tumors in the second highest dose of 2 mg/kg-day in this study. The
conclusion from the dose response analysis is that hexachlorobutadiene is carcinogenic only in the
kidney following ingestion at cytotoxic doses in the rat.
The nephrotoxicity of HCBD in rodents is dependent on a multi-step bioactivation
mechanism involving both kidney and liver enzymes. The ultimate step in this pathway is a P-lyase
mediated degradation of a HCBD metabolite that generates a highly reactive thioketene in proximal
tubule cells. In vitro studies suggest that cortical mitochondria are the critical subcellular target for
toxicity. Covalent binding of this reactive thioketene to cellular macromolecules (e.g. proteins,
HCBD February 2003 1-2
-------
mitochondrial DNA) and the resultant mitochondrial dysfunction is believed to underlie the
development of renal cytotoxicity and renal tubular adenomas and carcinomas in rodents.
Limited in vitro studies suggest humans have the ability to metabolize HCBD. The activity
of HCBD metabolizing enzymes, particularly renal p-lyase, may be many fold lower in humans than
the corresponding enzymes in rats (Lock, 1994; Lash et al., 1990; Anders and Dekant, 1998);
although the activity of glutathione S-transferase was not different between rodents and humans
(Dekant et al., 1998). Due to the fact that lower levels of reactive metabolites would be assumed
to form, there would be less concern for toxicity in humans.
The primary target organ for HCBD is the kidney. Individuals with preexisting kidney
damage may represent a potentially sensitive subpopulation for hexachlorobutadiene health effects.
Studies in animals showed that the young rats and mice were more sensitive to the acute effects of
oral HCBD than adults (Hook et al., 1983; Lock et al., 1984). Those data may suggest that infants
may potentially be more susceptible to hexachlorobutadiene toxicity.
Three key studies which have been used in determining points of departure for the reference
dose (RfD) are the Kociba et al. (1977) study and Schwetz et al. (1977) studies in rats and the NTP
(1991) study in mice. In the Kociba (1977) study, male and female Sprague-Dawley rats were fed
diets that contained 0, 0.2, 2, or 20 mg/kg-day HCBD for 22 months (males) or 24 months
(females). Non-neoplastic effects, including increased coproporphyrin excretion and microscopic
renal tubule epithelial hyperplasia, were seen at the two high doses, while renal tubule adenomas
and carcinomas were found only at the highest dose. In the Schwetz et al. (1977) study, male and
female Sprague-Dawley rats were fed a diet containing 0.2, 2.0, or 20 mg/kg-day HCBD for
evaluation of reproductive effects. HCBD was provided in the diet before and during mating, and
throughout gestation and lactation, for a total study duration of 148 days. At necropsy, relative
kidney weights were increased in high-dose males and females. Relative liver weight was increased
in high-dose males, and relative brain weight was increased in high-dose females. The kidneys of
males at the two high doses were roughened and had a mottled cortex. Histopathological
examination revealed dose-related increases in tubular dilation and regeneration in animals at the
two high doses. In the 13-week NTP (1991) study, B6C3FJ mice were fed a diet containing 0, 0.1,
0.4, 1.5, 4.9 or 16.8 mg/kg-day for males and 0, 0.2, 0.5, 1.8, 4.5 or 19.2 mg/kg-day for females.
Reduced body weight gain was reported in males at the two high doses and in females at the highest
dose. Relative kidney weight was decreased in the three high dose groups for males, and in females
in the highest dose group. High-dose males also exhibited decreased relative heart weight. Necropsy
revealed treatment-related increases in renal tubular cell regeneration in all dose groups in female
mice and in the two high dose groups in male mice.
A previous RfD for hexachlorobutadiene was 2 x 10"4 mg/kg-day (EPA, 1998a). It was
derived from a NOAEL of 0.2 mg/kg-day for renal tubular epithelial cell degeneration and
regeneration from the Kociba et al. (1977) study on rats and supported by the NTP (1991) study on
mice. A composite uncertainty factor (UF) of 1,000 was used in the derivation of the RfD.
The RfD for hexachlorobutadiene is 3 x 10"4 mg/kg-day. This is derived from the BMDL
of 0.1 mg/kg-day calculated from a BMD analysis of renal tubular epithelial cell regeneration from
the NTP (1991) study, with support from the Kociba (1977) and Schwetz (1977) studies in rats. A
HCBD February 2003 1-3
-------
composite UF of 300 is now used in the derivation of the RfD. The composite uncertainty factor
includes a factor of 10 to account for extrapolation from animals to humans; a factor of 10 for
protection of sensitive subpopulation; and a factor of 3 for database deficiencies (lack of a 2-
generation reproductive study and developmental toxicity studies in only one species). In accordance
with EPA's 1986 Guidelines for Carcinogen Risk Assessment (U. S. EPA, 1986), HCBD is classified
as a Group C (possible human) carcinogen. Under EPA's 1999 draft Guidelines for Carcinogen Risk
Assessment (U.S. EPA, 1999c), HCBD is classified as likely to be carcinogenic to humans by the
oral route of exposure. Its carcinogenic potential by the inhalation and dermal routes of exposure
is classified as cannot be determined because there are inadequate data to perform an assessment.
Two different approaches were used for dose-response extrapolation to estimate excess human
cancer risk for HCBD exposure from the rodent data. The default linearized multi-stage model
calculated a slope factor of 4 x 10"2 (mg/kg-day)"1, and a unit risk of 1.1 x 10"6 per |ig/L. For the
nonlinear approach, the RfD of 3 x IQ'4 mg/kg-day is used for the protection of cancer effect. In
consideration of the overall evidence, including the possible genotoxicity of HCBD metabolites, the
linear approach is recommended by EPA.
HCBD February 2003 1-4
-------
2.0 IDENTITY: PHYSICAL AND CHEMICAL PROPERTIES
The chemical and physical properties of hexachlorobutadiene (HCBD) are summarized in
Table 2-1. Synonyms for this chemical include perchlorobutadiene; l,l,2,3,4,4-hexachloro-l,3-
butadiene; 1,3-hexachlorobutadiene; Dolen-Pur; and GP-40-66:120.
HCBD is a colorless liquid at room temperature with a mild turpentine-like odor (HSDB,
2000). HCBD is poorly soluble in water, but is miscible in ethanol and ether (HSDB, 2000). HCBD
has a relatively low vapor pressure of 0.15 mm Hg at 25°C (U.S. EPA, 1991a). An odor threshold
of 0.006 mg/L has been reported for HCBD in water (U.S. EPA, 1980), and an air odor threshold
of 12 mg/m3 (Ruth, 1986). The Occupational Safety and Health Administration (OSHA)
Permissible Exposure Limit (PEL) is 0.21 mg/m3 (OSHA, 1989), making odor a poor warning
characteristic for HCBD. HCBD is characterized by high log Koc and log Kow values, 3.67 and 4.78,
respectively (ATSDR, 1994), reflecting properties which strongly influence its behavior
and fate in environmental media. The chemical structure of HCBD is shown in Figure 2-1.
Figure 2-1. Chemical Structure of Hexachlorobutadiene.
Cl {/ Cl
Cl
Hexachlorobutadiene
HCBD February 2003 1-1
-------
Table 2-1. Chemical and Physical Properties of Hexachlorobutadiene.
Property
Chemical Formula
Molecular Weight
Synonyms
Boiling Point (at 760 mm Hg)
Melting Point
Vapor Pressure (at 25°C)
Density (at 20°C)
Water Solubility (at 20°C)
Organic Solvents
Partition Coefficients
Odor Threshold (air)
Odor Threshold (water)
Conversion Factor
Information
C4C16
260.76
HCBD; Perchlorobutadiene;
Hexachlorbutadiene; 1, 1,2,3,4,4-Hexachloro-
1 , 3 -butadiene ; 1 , 3 -Hexachlorobutadiene ;
Dolen-Pur; GP-40-66:120NIOSH Registry of
Toxic Effects of Chemical Substances (RTECS)
No.EJ0700000U.S. EPA Hazardous Waste
No.U128Oil and Hazardous
Materials/Technical Assistance Data System
(OHM/TADS) No.OHM SlOOOllHazardous
Substances Data Bank (HSDB) No.2870
215°C
-21°C
0.15 mmHg
1.55 g/cm3
2-2.55mg/L
Ethanol, Ether
Log Kow 4.78
LogKoc3.67
12.00 mg/m3
0.006 mg/L
1 ppm = 10.66 mg/m3
1 mg/m3 = 0.0938 ppm
Sources: U.S. EPA (1980, 1991a); ChenJDplus (2000); HSDB (2000)
HCBD February 2003
2-2
-------
3.0 USES AND ENVIRONMENTAL FATE
3.1 Uses
HCBD has never been specifically manufactured as a commercial product in the United
States. However, significant quantities of the chemical are generated in the U.S. as waste by-product
from the chlorination of hydrocarbons such as tri- and tetrachloroethylene and carbon tetrachloride.
Lesser quantities have been imported in past decades for commercial purposes, mostly from
Germany. Until 1975 the most important commercial use of HCBD in the United States was for the
recovery of chlorine-containing gases in chlorine plants. Since then, HCBD has been used primarily
as a chemical intermediate in the production of rubber, and also as a hydraulic fluid, a fluid for
gyroscopes, a heat transfer liquid, a solvent, a laboratory reagent, a wash liquor for removing C4 and
higher hydrocarbons, and as a chemical intermediate in the production of chlorofluorocarbons and
lubricants (ATSDR, 1995; Howard, 1989). The chemical has also been used as a fumigant in
Russia, France, Italy, Greece, Spain, and Argentina, although use in the European Community is
reported to have ceased (van de Plassche and Schwegler, 2002).
Eight million pounds of HCBD were generated as a waste by-product in the U.S. in 1975,
with 0.1 million pounds released into the environment. By 1982, the annual U.S. by-product
generation of the chemical had increased to 28 million pounds. In contrast, the annual import rate
of HCBD dropped from 500,000 Ibs/yr imported annually in the late 1970s, to 145,000 Ibs/yr
imported in 1981 (ATSDR, 1994;Howard, 1989). Van de Plassche and Schwegler (2002) report that
all commercial production in Europe has ceased, and estimate that worldwide commercial
production has dropped from 10,000 tonnes in 1982 to virtually nil today.
3.2 Release to the Environment
HCBD is listed as a toxic release inventory (TRI) chemical. In 1986, the Emergency
Planning and Community Right-to-Know Act (EPCRA) established the Toxic Release Inventory
(TRI) of hazardous chemicals. Created under the Superfund Amendments and Reauthorization Act
(SARA) of 1986, EPCRA is also sometimes known as SARA Title III. The EPCRA mandates that
larger facilities publicly report when TRI chemicals are released into the environment. This public
reporting is required for facilities with more than 10 full-time employees that annually manufacture
or produce more than 25,000 pounds, or use more than 10,000 pounds, of TRI chemical (U.S. EPA,
1996c, 2000a).
Under these conditions, facilities are required to report the pounds per year of HCBD
released into the environment both on- and off-site. The on-site quantity is subdivided into air
emissions, surface water discharges, underground injections, and releases to land (see Table 3-1).
For HCBD, air emissions constitute most of the on-site releases. Also, over the period for which data
are available (1988-1998), surface water discharges generally increased, peaked in 1992-1993, and
then decreased significantly through the late 1990s. The TRI data for HCBD were reported from
eight States (CA, IL, KS, LA, NJ, NY, TX, UT); however, HCBD contamination has often been
found in remote areas far from apparent physical discharge sources (U.S. EPA, 2000b; Howard,
1989).
HCBD February 2003 3-1
-------
Table 3-1. Environmental Releases (in pounds) for Hexachlorobutadiene in the United
States, 1988-1998.
Year
1998
1997
1996
1995
1994
1993
1992
1991
1990
1989
1988
On-Site Releases
Air
Emissions
2,421
1,415
2,381
3,310
1,410
1,747
4,134
3,410
4,906
4,628
2,508
Surface Water
Discharges
5
9
256
661
351
1,200
1,911
681
715
622
153
Underground
Injection
0
299
952
434
201
520
738
200
330
330
220
Releases
to Land
0
0
0
0
0
0
0
2
0
1
0
Off-Site
Releases
510
200
310
252
430
12
5
4,263
45
26,343
19,640
Total On- &
Off-site
Releases
2,936
1,923
3,899
4,657
2,392
3,479
6,788
8,556
5,996
31,924
22,521
source: U.S. EPA (2000b)
Although the TRI data can be useful in giving a general idea of release trends, it is far from
exhaustive and has significant limitations. For example, only industries which meet TRI criteria (at
least 10 full-time employees and manufacture and processing of quantities exceeding 25,000 Ibs/yr,
or use of more than 10,000 Ibs/yr) are required to report releases. These reporting criteria do not
account for releases from smaller industries. Threshold manufacture and processing quantities also
changed from 1988-1990 (dropping from 75,000 Ibs/yr in 1988 to 50,000 Ibs/yr in 1989 to its
current 25,000 Ibs/yr in 1990) creating possibly misleading data trends. Finally, the TRI data is
meant to reflect releases and should not be used to estimate general exposure to a chemical (U.S.
EPA, 2000c,d).
While TRI releases were reported in only eight States, the use of HCBD is widespread. It is
included in the Agency for Toxic Substances and Disease Registry' s (ATSDR) Hazardous Substance
Release and Health Effects Database (HazDat) and has been detected in site samples in fourteen
States (AL, AZ, CT, IA, LA, MI, MN, NJ, NY, OH, PA, RI, SC, WA; ATSDR, 2000). These States
are distributed nationwide and include 11 States and two regions (New England and the Pacific
Northwest) not reporting TRI releases yet manifesting HCBD detections in the environment.
The National Priorities List (NPL) of hazardous waste sites, created in 1980 by the
Comprehensive Environmental Response, Compensation and Liability Act (CERCLA), is a listing
of some of the most health-threatening waste sites in the United States. HCBD was detected in
el even NPL sites in 1999. These sites are located in eight States: AK, CO, IN, LA, NJ, OH, PA, WA.
Again, note that there is little overlap between these States and the eight TRI reporting States (U.S.
EPA, 1999a).
In summary, although HCBD is not manufactured in the United States, both its use in
industry and occurrence in the environment are widespread. Significant quantities of HCBD are
generated in the United States as a waste by-product, and smaller quantities are imported for
industrial needs. HCBD is present in hazardous waste sites in at least 8 States (at NPL sites), has
HCBD February 2003
3-2
-------
been detected in site samples in at least 14 States (listed in ATSDR' s HazDat), and has been released
into the environment directly in at least 8 States (based on TRI data).
3.3 Fate in Air
HCBD is released to air via chemical manufacturing and processing and by waste
incineration (HSDB, 2000). Modeling and monitoring data suggest that the atmospheric burden of
HCBD in the northern hemisphere is approximately 3.2 million kg/yr (Class and Ballschmiter,
1987). Dispersion of HCBD in the atmosphere has been confirmed by detection of HCBD at
locations distant from sources of release (WHO, 1994). The high log organic carbon partition
coefficient (log Koc) of HCBD indicates that it will readily adsorb to airborne parti culate matter with
a high organic content. Thus, HCBD in air is found both as a vapor and in association with
atmospheric particulates.
No specific information is available on the transformation and degradation of HCBD in air.
3.4 Fate in Water
HCBD is released to surface and ground water via industrial effluents, by leaching from
landfills or soil, or by urban runoff (ATSDR, 1994). Sorption to sediments and suspended
particulates is an important factor in the fate of HCBD in water (U.S. EPA, 199la). As a result of
this affinity for particulates and sediments, HCBD-contaminated areas will usually have higher
sediment concentrations than water concentrations of the chemical. U.S. EPA (1976) found that
HCBD concentrations in the Mississippi delta water were <2 |ig/L, while concentrations in mud or
soil were >200 |ig/L. Leeuwangh et al. (1975) observed that equilibration of initially
uncontaminated sediment with HCBD-contaminated water resulted in sediment concentrations 100-
fold greater than those observed in the water.
Volatilization of HCBD from water to air also occurs, although the low vapor pressure of
HCBD (0.15 mmHg at 25°C) suggests that this process may occur relatively slowly (U.S. EPA,
199 la). Limited data are available on the transformation and degradation of HCBD in water. Under
aerobic conditions in batch culture, complete biodegradation has been observed to occur in sewage-
inoculated waters after seven days (Tabak et al., 1981). These data suggest that HCBD may
biodegrade in natural waters. In contrast, no degradation was observed under anaerobic conditions
in a separate study (Johnson and Young, 1983). No data were available on hydrolysis or photolysis
of HCBD in water. Estimates of HCBD half-life range from 3 to 30 days in rivers and 30 to 300 days
in lakes and groundwater (Zoeteman et al., 1980).
The high octanol-water partition coefficient (Kow) of HCBD suggests that this chemical can
readily partition from water into biota. Laboratory and field investigations confirm that HCBD has
bioaccumulation potential (WHO, 1994). Field-measured bioaccumulation factors range from 46
to 27,780 (U.S. EPA, 1999b). No evidence for biomagnification has been observed in laboratory or
field studies (WHO, 1994).
HCBD February 2003 3-3
-------
3.5 Fate in Soil
HCBD can be released to soil by disposal of industrial waste in landfill operations (ATSDR,
1994). Volatilization from soil surfaces is expected to be a primary process for loss of HCBD from
soil (Tabak et al., 1981). However, as HCBD readily adsorbs to soil organic particles, volatilization
from highly organic soils is predicted to be low (HSDB, 2000).
No data regarding transformation or degradation of HCBD in soil were located. Data from
experiments conducted in water (Tabak et al., 1981) suggest that biodegradation will occur if aerobic
conditions are present (HSDB, 2000). Results obtained in sludge incubated under anaerobic
conditions (Johnson and Young, 1983) suggest that biodegradation will not occur under anaerobic
soil conditions. Soil organic matter content is likely to be an important factor in biodegradation time,
since adsorption of HCBD to organic matter will significantly decrease its bioavailability to
microorganisms. In the absence of significant biodegradation or other loss processes, persistence of
HCBD in soil may allow migration of the compound into groundwater, particularly in sandy soils
(U.S. EPA, 1984).
HCBD February 2003 3-4
-------
4.0 EXPOSURE FROM DRINKING WATER
This section of the report examines the occurrence of HCBD in drinking water. While no
complete national database exists of unregulated or regulated contaminants in drinking water from
public water systems (PWSs) collected under SDWA, this report aggregates and analyzes existing
State data that have been screened for quality, completeness, and representativeness. Populations
served by PWSs exposed to HCBD are estimated, and the occurrence data are examined for regional
or other special trends. To augment the incomplete national drinking water data and aid in the
evaluation of occurrence, information on ambient occurrence of HCBD is also reviewed.
4.1 Ambient Occurrence
To understand the presence of a chemical in the environment, an examination of ambient
occurrence is useful. In a drinking water context, ambient water is source water existing in surface
waters and aquifers before treatment. The most comprehensive and nationally representative data
describing ambient water quality in the United States are being produced through the United States
Geological Survey's (USGS) National Water Quality Assessment (NAWQA) program. NAWQA,
however, is a relatively young program and complete national data are not yet available from their
entire array of sites across the nation.
4.1.1 Data Sources and Methods
To examine water quality status and trends in the United States, the USGS instituted the
NAWQA program in 1991. NAWQA is designed and implemented in such a manner to allow
consistency and comparison between representative study basins located around the country,
facilitating interpretation of natural and anthropogenic factors affecting water quality (Leahy and
Thompson, 1994).
The NAWQA program consists of 59 significant watersheds and aquifers referred to as
"study units." The study units represent approximately two thirds of the overall water usage in the
United States and a similar proportion of the population served by public water systems.
Approximately one half of the Nation's land area is represented (Leahy and Thompson, 1994).
To facilitate management and make the program cost-effective, approximately one third of
the study units at a time engage in intensive assessment for a period of 3 to 5 years. This is followed
by a period of less intensive research and monitoring that lasts between 5 and 7 years. This way all
59 study units rotate through intensive assessment over a ten-year period (Leahy and Thompson,
1994). The first round of intensive monitoring (1991-1996) targeted 20 watersheds. Thi s first group
was more heavily slanted toward agricultural basins. A national synthesis of results from these study
units and other research initiatives focusing on pesticides and nutrients is being compiled and
analyzed (Kolpin et al., 1998; Larson et al., 1999).
For volatile organic chemicals (VOCs), the national synthesis will compile data from the first
and second rounds of intensive assessments. Study units assessed in the second round represent
conditions in more urbanized basins, but initial results are not yet available. However, VOCs were
analyzed in the first round of intensive monitoring and data are available for these study units
HCBD February 2003 4-1
-------
(Squillace et al., 1999). The minimum reporting limit (MRL) for most VOCs, including HCBD, was
0.2 |ig/L (Squillace etal., 1999). Additional information on analytical methods used in the NAWQA
study units, including method detection limits, are described by Gilliom and others (in press).
Furthermore, the NAWQA program has compiled, by study unit, data collected from local,
State, and other Federal agencies to augment its own data. The data set provides an assessment of
VOCs in untreated ambient groundwater of the conterminous United States for the period
1985-1995 (Squillace et al., 1999). Data were included in the compilation if they met certain criteria
for collection, analysis, well network design, and well construction (Lapham et al., 1997). They
represent both rural and urban areas, but should be viewed as a progress report as NAWQA data
continue to be collected that may influence conclusions regarding occurrence and distribution of
VOCs (Squillace et al., 1999).
4.1.2 Results
Initial results published for the 20 NAWQA study units undergoing intensive assessment
from 1991-1996 indicate that HCBD was not detected in ground water (Squillace et al., 1999).
HCBD also was not detected in rural or urban wells of the local, State, and federal data set compiled
by NAWQA. These data represent untreated ambient ground water of the conterminous United
States for the years 1985-1995 (Squillace et al., 1999).
Furthermore, a review of highway and urban runoff studies found no detections of HCBD
(Lopes and Dionne, 1998). This review was undertaken as part of the National Highway Runoff
Data and Methodology Synthesis and examined 44 studies implemented since 1970.
4.2 Drinking Water Occurrence
The Safe Drinking Water Act (SOWA), as amended in 1986, required Public Water Systems
(PWSs) to monitor for specified "unregulated" contaminants, on a five year cycle, and to report the
monitoring results to the States. Unregulated contaminants do not have an established or proposed
National Primary Drinking Water Regulation (NPDWR), but they are contaminants that were
formally listed and required for monitoring under federal regulations. The intent was to gather
scientific information on the occurrence of these contaminants to enable a decision as to whether
or not regulations were needed. All non-purchased community water systems (CWSs) and non-
purchased non-transient non-community water systems (NTNCWSs), with greater than 150 service
connections, were required to conduct this unregulated contaminant monitoring. Smaller systems
were not required to conduct this monitoring under federal regulations, but were required to be
available to monitor if the State decided such monitoring was necessary. Many States collected data
from smaller systems. Additional contaminants were added to the Unregulated Contaminant
Monitoring (UCM) program in 1991 (U.S. EPA, 199Ic) for required monitoring that began in 1993
[57 FR 31776] (U.S. EPA, 1992c).
HCBD has been monitored under the SDWA UCM program since 1987 [52 FR 25720].
Monitoring for HCBD under UCM continued throughout the 1990s, but ceased for small public
water systems (PWSs) under a direct final rule published on January 8, 1999 (64 FR 1494).
Monitoring ended for large PWSs with promulgation of the new Unregulated Contaminant
HCBD February 2003 4-2
-------
Monitoring Regulation (UCMR) issued September 17, 1999 (64 FR 50556) and effective January
1, 2001. At the time the UCMR lists were developed, the Agency concluded there were adequate
monitoring data for a regulatory determination. This obviated the need for continued monitoring
under the new UCMR list.
4.2.1 Data Sources, Data Quality, and Analytical Methods
Currently, there is no complete national record of unregulated or regulated contaminants in
drinking water from public water systems collected under SDWA. Many States have submitted their
unregulated contaminant PWS monitoring data to EPA databases, but there are issues of data
quality, completeness, and representativeness. Nonetheless, a significant amount of State data are
available for UCM contaminants that can provide estimates of national occurrence.
The National Contaminant Occurrence Database (NCOD) is an interface to the actual
occurrence data stored in a database called the Safe Drinking Water Information System (Federal
version; SDWIS/FED) and can be queried to provide a summary of the data in SDWIS/FED for a
particular contaminant. The data used in this report were derived from the data in SDWIS/FED and
another database called the Unregulated Contaminant Information System (URCIS).
The data in this report have been reviewed, edited, and filtered to meet various data quality
objectives for the purposes of this analysis. Hence, not all data from a particular source were used,
only data meeting the quality objectives described below. The sources of these data, their quality
and national aggregation, and the analytical methods used to estimate a given contaminant's national
occurrence (from these data) are discussed in this section (for further details see U.S. EPA,
2001a,b).
UCM Rounds 1 and 2
The 1987 UCM contaminants include 34 volatile organic compounds (VOCs), divided into
two groups: one with 20 VOCs for mandatory monitoring, and the other with 14 VOCs for
discretionary monitoring [52FR25720]. HCBD was among the 14 VOCs included for discretionary
monitoring. The UCM (1987) contaminants were first monitored coincident with the Phase I
regulated contaminants, during the 1988-1992 period. This period is often referred to as "Round 1"
monitoring. The monitoring data collected by the PWSs were reported to the States (as primacy
agents), but there was no protocol in place to report these data to EPA. These data from Round 1
were collected by EPA from many States over time.
The Round 1 data were collected in the URCIS. Most of the Phase 1 regulated contaminants
were also VOCs. Both unregulated and regulated VOCs are analyzed using the same sample and the
same laboratory methods. Hence, the URCIS database includes data on all of these 62 contaminants:
the 34 UCM (1987) VOCs; the 21 regulated Phase 1 VOCs; 2 regulated synthetic organic
contaminants (SOCs); and 5 miscellaneous contaminants that were voluntarily reported by some
States (e.g., isomers of other organic contaminants).
The 1993 UCM contaminants include 13 SOCs and 1 inorganic contaminant (IOC) [56 FR
3 526]. Monitoring for the UCM (1993) contaminants began coincident with the Phase II/V regulated
HCBD February 2003 4-3
-------
contaminants in 1993 through 1998. This is often referred to as "Round 2" monitoring. The UCM
(1987) contaminants were also included in the Round 2 monitoring. As with other monitoring data,
PWSs reported these results to the States. EPA, during the past several years, requested that the
States submit these historic data to EPA.
The details of the actual individual monitoring periods are complex. The timing of required
monitoring was staggered related to different size classes of PWSs, and the program was
implemented somewhat differently by different States. Round 1 includes the period from
1988-1992, it also includes results from samples analyzed prior to 1988 (for further details see U.S.
EPA, 2001a,b).
Developing a Nationally Representative Perspective
The URCIS and SDWIS/FED databases contain contaminant occurrence data from a total
of 40 and 35 primacy entities (largely States), respectively. However, data from some States are
incomplete and biased. Furthermore, the national representativeness of the data is questionable
because the data were not collected in a systematic or random statistical framework. These State data
could be heavily skewed to low-occurrence or high-occurrence settings. Hence, the data were
evaluated based on pollution-potential indicators and the spatial/hydrologic diversity of the nation.
This evaluation enabled the construction of a cross-section from the available State data sets that
provides a reasonable representation of national occurrence.
A national cross-section from State SDWA contaminant databases was established using the
approach developed for the EPA report A Review of Contaminant Occurrence in Public Water
Systems (U. S. EPA, 1999c). This approach was developed to support occurrence analyses for EPA's
Chemical Monitoring Reform (CMR) evaluation. It was supported by peer reviewers and
stakeholders because it is clear, repeatable, and understandable. The approach cannot provide a
"statistically representative" sample because the original monitoring data were not collected or
reported in an appropriate fashion. However, the resultant "national cross-section" of States should
provide a clear indication of the central tendency of the national data. The remainder of this section
provides a summary description of how the national cross-sections for the URCIS (Round 1) and
SDWIS/FED (Round 2) databases were developed. The details of the approach are presented in U. S.
EPA(2001a,b).
Cross-Section Development
As a first step in developing the cross-section, the State data contained in the URCIS
database (which contains Round 1 monitoring results) and SDWIS/FED database (which contains
Round 2 monitoring results) were evaluated for completeness and quality. For both the URCIS
(Round 1) and SDWIS/FED (Round 2) databases, some State data were unusable for a variety of
reasons. Some States reported only detections, or their data had incorrect units. Datasets only
including detections are obviously biased. Other problems included incomplete data sets without all
PWSs reporting. Also, data from Washington, D.C. and the Virgin Islands were excluded from this
analysis because it was difficult to evaluate them for the current purposes in relation to complete
State data (U.S. EPA, 2001a, Sections II and III).
HCBD February 2003 4-4
-------
The balance of the States remaining after the data quality screening were then examined to
establish a national cross-section. This step was based on evaluating the States' pollution potential
and geographic coverage in relation to all States. Pollution potential is considered to ensure a
selection of States that represent the range of likely contaminant occurrence and a balance with
regard to likely high and low occurrence. Geographic consideration is included so that the wide
range of climatic and hydrogeologic conditions across the United States are represented, again
balancing the varied conditions that affect transport and fate of contaminants (U.S. EPA, 2001b,
Sections III.A. and III.B.).
The cross-section States were selected to represent a variety of pollution potential conditions.
Two primary pollution potential indicators were used. The first factor selected indicates pollution
potential from manufacturing/population density and serves as an indicator of the potential for VOC
contamination within a State. Agriculture was selected as the second pollution potential indicator
because the majority of SOCs of concern are pesticides (U.S. EPA, 200 Ib, Section III. A.). The 50
individual States were ranked from highest to lowest based on the pollution potential indicator data.
For example, the State with the highest ranking for pollution potential from manufacturing received
a ranking of 1 for this factor and the State with the lowest value was ranked as number 50. States
were ranked for their agricultural chemical use status in a similar fashion.
The States' pollution potential rankings for each factor were subdivided into four quartiles
(from highest to lowest pollution potential). The cross-section States were chosen from all quartiles
for both pollution potential factors to ensure representation, as follows: States with high
agrochemical pollution potential rankings and high manufacturing pollution potential rankings;
States with high agrochemical pollution potential rankings and low manufacturing pollution
potential rankings; States with low agrochemical pollution potential rankings and high
manufacturing pollution potential rankings; and States with low agrochemical pollution potential
rankings and low manufacturing pollution potential rankings ( U.S. EPA, 200Ib, Section III.B.). In
addition, some secondary pollution potential indicators were considered to further ensure that the
cross-section States included the spectrum of pollution potential conditions (high to low).
The data quality screening, pollution potential rankings, and geographic coverage analysis
established national cross-sections of 24 Round 1 (URCIS) States and 20 Round 2 (SDWIS/FED)
States. In each cross-section, the States provide good representation of the Nation's varied climatic
and hydrogeologic regimes and the breadth of pollution potential for the contaminant groups (Table
4-1 and Figure 4-1).
Cross-Section Evaluation
To evaluate and validate the method for creating the national cross-sections, the method was
used to create smaller State subsets from the 24-State, Round 1 cross-section and aggregations.
Again, States were chosen to achieve a balance from the quartiles describing pollution potential, and
a balanced geographic distribution, to incrementally build
HCBD February 2003 4-5
-------
Table 4-1. Cross-section States for Round 1 (24 States) and Round 2 (20 States).
Round 1 (URCIS)
Alabama
Alaska*
Arizona
California
Florida
Georgia
Hawaii
Illinois
Indiana
Iowa
Kentucky*
Maryland*
Minnesota*
Montana
New Jersey
New Mexico*
North Carolina*
Ohio*
South Dakota
Tennessee
Utah
Washington*
West Virginia
Wyoming
Round 2 (SDWIS/FED)
Alaska*
Arkansas
Colorado
Kentucky*
Maine
Maryland*
Massachusetts
Michigan
Minnesota*
Missouri
New Hampshire
New Mexico*
North Carolina*
North Dakota
Ohio*
Oklahoma
Oregon
Rhode Island
Texas
Washington*
* cross-section State in both Round 1 and Round 2
Figure 4-1. Geographic Distribution of Cross-section States for Round 1 (left) and Round
2 (right).
subset cross-sections of various sizes. For example, the Round 1 cross-section was tested with
subsets of 4, 8 (the first 4-State subset plus 4 more States), and 13 (8-State subset plus 5) States.
Two additional cross-sections were included in the analysis for comparison: a cross-section
composed of the 16 biased States eliminated from the 24-State cross-section for data quality reasons
and a cross-section composed of all 40 Round 1 States (U.S. EPA, 2001, Section III.B.I).
These Round 1 incremental cross-sections were then used to evaluate occurrence for an array
of both high and low occurrence contaminants. The comparative results illustrate several points. The
HCBD February 2003
4-6
-------
results are quite stable and consistent for the 8-, 13- and 24-State cross-sections. They are much less
so for the 4-State, 16-State (biased), and 40-State (all Round 1 States) cross-sections. The 4-State
cross-section is apparently too small to provide balance both geographically and with pollution
potential, a finding that concurs with past work (U.S. EPA, 1999c). The CMR analysis suggested
that a minimum of 6-7 States was needed to provide balance both geographically and with pollution
potential, and the CMR report used 8 States out of the available data for its nationally representative
cross-section. The 16-State and 40-State cross-sections, both including the biased States, provided
occurrence results that were unstable and inconsistent for a variety of reasons associated with their
data quality problems ( U.S. EPA, 2001, Section III.B.l).
The 8-, 13-, and 24-State cross-sections provide very comparable results, are consistent, and
are usable as national cross-sections to provide estimates of contaminant occurrence. Including
greater data from more States improves the national representation and the confidence in the results,
as long as the States are balanced related to pollution potential and spatial coverage. The 24- and
20-State cross-sections provide the best, nationally representative cross-sections for the Round 1 and
Round 2 data.
Data Management and Analysis
The cross-section analyses focused on occurrence at the water system level; i.e., the
summary data presented discuss the percentage of public water systems with detections, not the
percentage of samples with detections. By normalizing the analytical data to the system level,
skewness inherent in the sample data, particularly over the multi-year period covered in the URCIS
data, is avoided. System level analysis was used since a PWS with a known contaminant problem
usually has to sample more frequently than a PWS that has never detected the contaminant.
Obviously, the results of a simple computation of the percentage of samples with detections (or other
statistics) can be skewed by the more frequent sampling results reported by the contaminated site.
This level of analysis is conservative. For example, a system need only have a single sample with
an analytical result greater than the MRL, i.e., a detection, to be counted as a system with a result
"greater than the MRL."
Also, the data used in the analyses were limited to only those data with confirmed water
source and sampling type information. Only standard SDWA compliance samples were used;
"special" samples, or "investigation" samples (investigating a contaminant problem that would bias
results), or samples of unknown type were not used in the analyses. Various quality control and
review checks were made of the results, including follow-up questions to the States providing the
data. Many of the most intractable data quality problems encountered occurred with older data.
These problematic data were, in some cases, simply eliminated from the analysis. For example,
when the number of data with problems were insignificant relative to the total number of
observations, they were dropped from the analysis (For further details, see Cadmus, 2000).
Occurrence Analysis
To evaluate national contaminant occurrence, a two-stage analytical approach has been
developed. The first stage of analysis provides a straight-forward, conservative, broad evaluation
of occurrence of the Contaminant Candidate List (CCL) preliminary regulatory determination
HCBD February 2003 4-7
-------
priority contaminants as described above. These descriptive statistics are summarized here. Based
on the findings of the Stage 1 Analysis, EPA will determine whether more intensive statistical
evaluations, the Stage 2 Analysis, may be warranted to generate national probability estimates of
contaminant occurrence and exposure for priority contaminants (for details on this two stage
analytical approach see Cadmus, 2000)
The summary descriptive statistics presented in Table 4-2 for HCBD are a result of the Stage
1 analysis and include data from both Round 1 (URCIS, 1987-1992) and Round 2 (SDWIS/FED,
1993-1997) cross-section States. Included are the total number of samples, the percent samples with
detections, the 99th percentile concentration of all samples, the 99th percentile concentration of
samples with detections, and the median concentration of samples with detections. The percentages
of PWSs and population served indicate the proportion of PWSs whose analytical results showed
a detection(s) of the contaminant (simple detection, > MRL) at any time during the monitoring
period; or a detection(s) greater than half the Health Reference Level (HRL); or a detection(s)
greater than the Health Reference Level. The Health Reference Level, 0.9 |ig/L, is a preliminary
estimated health effect level used for this analysis. The HRL was derived using the 10"6 cancer risk
as calculated by the linear method using a body weight to the three quarter power (section 8.8.2;
slope factor 4 x 10"2 (mg/kg/day)"1.
When monitoring results were compared to a value of one-half of the HRL, 0.16% of Round
1 (106 systems) and 0.08% of Round 2 (51 systems) water supplies exceeded this benchmark at least
once during the reporting period. The percentages of water supplies that exceeded the HRL at least
once in Round 1 and Round 2 monitoring were 0.11% (74 systems) and 0.02% (11 systems),
respectively.
The 99th percentile concentration is used here as a summary statistic to indicate the upper
bound of occurrence values because maximum values can be extreme values (outliers) that
sometimes result from sampling or reporting error. The 99th percentile concentration is presented for
both the samples with only detections and all of the samples because the value for the 99th percentile
concentration of all samples is below the MRL (denoted by "<" in Table 4-2). For the same reason,
summary statistics such as the 95th percentile concentration of all samples or the median (or mean)
concentration of all samples are omitted because these also are all "<" values. This is the case
because only 0.1 to 0.05% of all samples recorded detections of HCBD in Round 1 and Round 2.
As a convention, a value of half the MRL is often used as an estimate of the concentration
of a contaminant in samples/systems whose results are less than the MRL. With a contaminant with
relatively low occurrence such as HCBD in drinking water occurrence databases, the median or
mean value of occurrence using this assumption would be half the MRL (0.5 x MRL). However, for
these occurrence data this is not straightforward. For Round 1 and Round 2, States have reported a
wide range of values for the MRLs. This is in part related to State data management differences as
well as real differences in analytical methods, laboratories, and other factors.
The situation can cause confusion when examining descriptive statistics for occurrence. For
example, the modal MRL value for the Round 1 samples is 0.50 |ig/La value twice as large as the
median concentration of detections for Round 1 (0.25 |ig/L) (This occurs because some States and/or
systems reporting detections were using a lower MRL and had positive results lower than the MRL
HCBD February 2003 4-8
-------
used by other States or systems). For Round 2, most States reported non-detections as zeros resulting
in a modal MRL value of zero. By definition the MRL cannot be zero. This is an artifact of State
data management systems. Because a simple meaningful summary statistic is not available to
describe the various reported MRLs, and to avoid confusion, MRLs are not reported in the summary
table, but rather are designated as "variable" (Table 4-2).
In Table 4-2, national occurrence is estimated by extrapolating the summary statistics for the
24- and 20-State cross-sections to national numbers for systems, and population served by systems,
from the Water Industry Baseline Handbook, Second Edition (U.S. EPA, 2000e). From the
handbook, the total number of community water systems (CWSs) plus non-transient, non-
community water systems (NTNCWSs) is 65,030, and the total population served by CWSs plus
NTNCWSs is 213,008,182 persons (see Table 4-2). To arrive at the national occurrence estimate for
a particular cross-section, the national estimate for PWSs (or population served by PWSs) is simply
multiplied by the percentage for the given summary statistic, [i.e., for Round 1, the national estimate
for the total number of PWSs with detections (228) is the product of the percentage of Round 1
PWSs with detections (0.35%) and the national estimate for the total number of PWSs (65,030)].
Because the State data used for the cross-section are not a strict statistical sample, national
extrapolations of these Stage 1 analytical results can be problematic, especially for contaminants
with very low occurrence like hexachlorobutadiene and other CCL regulatory determination priority
contaminants. For this reason, the nationally extrapolated estimates of occurrence based on Stage
1 results are not presented in the CCL Federal Register Notice. The presentation in the Federal
Register Notice of only the actual results of the cross-section analysis maintains a straight-forward
presentation, and the integrity of the data, for stakeholder review. The nationally extrapolated Stage
1 occurrence values are presented here, however, to provide additional perspective. A more rigorous
statistical modeling effort, the Stage 2 analysis, could be conducted on the cross-section data
(Cadmus, 2001). The Stage 2 results would be more statistically robust and more suitable to national
extrapolation. This approach would provide a probability estimate and would also allow for better
quantification of estimation error.
Round 1(1987-1992) and Round 2 (1993-1997) data were not merged because they
represent different time periods, different States (only eight States are represented in both rounds),
and each round has different data management and data quality problems. The two rounds are only
merged for the simple spatial analysis overview presented in Section 4.2 and Figures 4-2 and 4-4.
4.2.2 Results
Occurrence Estimates
While States with detections of HCBD are widespread (Figure 4-2), the percentages of PWSs
by State with detections are low (Table 4-2). In aggregate, the cross-sections show only 0.2-0.4 %
of PWSs in both rounds experienced detections (>MRL), affecting 0.9-2.4% of the population
served (approximately 2-5 million people). Percentages of PWSs with detections greater than half
the Health Reference Level (> V2 HRL) are slightly lower: 0.1-0.2%. The percentage of PWSs
exceeding the Health Reference Level (> HRL) for both rounds is very small (see also Figure 4-4).
HCBD February 2003 4-9
-------
The percentage of PWSs that experienced detections >HRL in Rounds 1 and 2 are 0.1% and 0.02%,
respectively; affecting a population of approximately 780,000 and 10,000, respectively.
There are some qualifying notes for both rounds of data that warrant discussion. The Round
1 estimates of PWSs affected by HCBD are influenced by the State of Florida (Table 4-2; Figures
4-3 and 4-4). This State reports that 5.4% of its PWSs experienced detections greater than the HRL
during Round 1, a value considerably greater than the next highest State (1.5%). This suggests that
Florida's data for HCBD is incomplete and may be biased. Out of 855 Florida PWSs reporting
contaminant data for Round 1 monitoring, only 112 provided data for HCBD (U.S. EPA, 200la).
Also, the 5.4% of systems reporting detections all reported concentrations greater than the Health
Reference Level. These figures suggest that perhaps only systems experiencing problems submitted
data for HCBD, biasing Florida's results for occurrence measures examined in this report.
The large values for the Round 2 national estimates of population served with detections
greater than the MRL and greater than half the HRL are influenced by the inclusion of one PWS
serving a very large population (1.5 million people). While the percentage of systems with
detections of HCBD are similar (both rounds show low values, 0.2-0.4% PWSs > MRL), the
difference in population served results in a larger difference in the population extrapolations.
Note that for the Round 1 cross-section, the total number of PWSs (and the total population
served by the PWSs) is not the sum of the number of ground water and surface water systems (or
the populations served by those systems). Because some public water systems are seasonally
classified as either surface or ground water, some systems may be counted in both categories. The
population numbers for the Round 1 cross-section are also incomplete. Not all of the PWSs for
which occurrence data was submitted reported the population they served. However, the population
numbers presented in Table 4-2 for the Round 1 cross-section are reported from 94% of the systems.
The national estimates extrapolated from Round 1 and Round 2 PWS, numbers and
populations are not additive. In addition to the Round 1 classification and reporting issues outlined
above, the proportions of surface water and ground water PWSs, and populations served by them,
are different between the Round 1 and 2 cross-sections and the national estimates. For
example, approximately 48% of the population served by PWSs in the Round 1 cross-section States
are served by surface water PWSs (Table 4-2). Nationally, however, that proportion changes to 60%.
HCBD February 2003 4-10
-------
Table 4-2. Summary Occurrence Statistics for Hexachlorobutadiene.
Frequency Factors
Total Number of Samples
Percent of Samples with Detections
99th Percentile Concentration (all samples)
Health Reference Level
Minimum Reporting Level (MRL)
99th Percentile Concentration of Detections
Median Concentration of Detections
Total Number of PWSs
Number of GW PWSs
Number of SW PWSs
Total Population
Population of GW PWSs
Population of SW PWSs
24-State
Cross-Section1
(Round 1)
42,839
0.13%
< (Non-detect)
0.9 |ig/L
Variable*
10|ig/L
0.25 |ig/L
12,284
10,980
1,385
71,582,571
40,399,177
34,418,834
20-State
Cross-Section2
(Round 2)
93,585
0.05%
< (Non-detect)
0.9 |ig/L
Variable*
1.5 |ig/L
0.30 |ig/L
22,736
20,380
2,356
67,075,493
24,960,222
42,115,271
Occurrence by System
PWSs with detections (> MRL)
Range of Cross-Section States
GW PWSs with detections
SW PWSs with detections
PWSs > !/2 Health Reference Level (HRL)
Range of Cross-Section States
GW PWSs > !/2 Health Reference Level
SW PWSs > !/2 Health Reference Level
PWSs > Health Reference Level
Range of Cross-Section States
GW PWSs > Health Reference Level
SW PWSs > Health Reference Level
Occurrence by Population Served
PWS Population Served with detections
0.350%
0-5.36%
0.301%
0.722%
0.163%
0-5.36%
0.118%
0.505%
0.114%
0-5.36%
0.064%
0.505%
0.896%
0.180%
0-3.36%
0.132%
0.594%
0.079%
0-0.51%
0.064%
0.212%
0.018%
0- 0.24%
0.005%
0.127%
2.360%
National System &
Population Numbers3
-
-
-
-
-
-
-
65,030
59,440
5,590
213,008,182
85,681,696
127,326,486
National Extrapolation4
Round 1
228
N/A
179
40
106
N/A
70
28
74
N/A
38
28
Round 1
1,909,000
Round 2
117
N/A
79
33
51
N/A
38
12
11
N/A
3
7
Round 2
5,027,000
HCBD February 2003
4-11
-------
Table 4-2 (continued)
Frequency Factors
Range of Cross-Section States
GW PWS Population with detections
SW PWS Population with detections
PWS Population Served > !/2 Health Ref Level
Range of Cross-Section States
GW PWS Population > !/2 Health Ref Level
SW PWS Population > !/2 Health Ref Level
PWS Population Served > Health Ref Level
Range of Cross-Section States
GW PWS Population > Health Ref Level
SW PWS Population > Health Ref Level
24-State
Cross-Section1
(Round 1)
0-11.38%
1.458%
0.153%
0.569%
0-11.38%
0.978%
0.036%
0.367%
0- 9.66%
0.619%
0.036%
20-State
Cross-Section2
(Round 2)
0-29.93%
0.186%
3.649%
2.331%
0-29.92%
0.177%
3.607%
0.005%
0- 0.02%
0.011%
0.001%
National System &
Population Numbers3
N/A
1,249,000
194,000
1,213,000
N/A
838,000
46,000
781,000
N/A
531,000
46,000
N/A
159,000
4,646,000
4,965,000
N/A
152,000
4,593,000
10,000
N/A
9,000
1,000
1. Summary Results based on data from 24-State Cross-Section, from URCIS, UCM (1987) Round 1.
2. Summary Results based on data from 20-State Cross-Section, from SDWIS/FED, UCM (1993) Round 2.
3. Total PWS and population numbers are from EPA March 2000 Water Industry Baseline Handbook.
4. National extrapolations are from the 24-State data and 20-State data using the Baseline Handbook system and population numbers.
* see text for discussion
- PWS = Public Water Systems; GW = Ground Water; SW = Surface Water; MRL = Minimum Reporting Level (for laboratory analyses);
- Health Reference Level = Health Reference Level, an estimated health effect level used for preliminary assessment for this review; N/A = Not
Applicable
- The Health Reference Level used for hexachlorobutadiene is 0.9 (ig/L. This is a draft value for working review only.
- Total Number of Samples = the total number of analytical records for hexachlorobutadiene.
- 99th Percentile Concentration = the concentration value of the 99th percentile of either all analytical results or just the samples with detections (in
- Median Concentration of Detections = the median analytical value of all the detections (analytical results greater than the MRL) (in (ig/L).
- Total Number of PWSs = the total number of public water systems with records for hexachlorobutadiene.
- Total Population Served = the total population served by public water systems with records for hexachlorobutadiene.
- % PWS with detections, % PWS > '/2 Health Reference Level, % PWS > Health Reference Level = percent of the total number of public water
systems with at least one analytical result that exceeded the MRL, Vt Health Reference Level, Health Reference Level, respectively.
- % PWS Population Served with detections, % PWS Population Served >'/2 Health Reference Level, % PWS Population Served > Health Reference
Level = percent of the total population served by PWSs with at least one analytical result exceeding the MRL, 1A Health Reference Level, or the Health
Reference Level, respectively.
Both Round 1 and Round 2 national cross-sections show a proportionate balance in source
waters. Nationally, 91% of PWSs use ground water (and 9% surface waters): Round 1 shows 89%,
and Round 2 shows 90% of systems using ground water. The relative populations served are not as
closely comparable. Nationally, about 40% of the population is served by PWSs using groundwater
(and 60% by surface water). Round 2 data is most representative with 37% of the cross-section
population served by ground water; Round 1 shows about 55%.
There are differences in the occurrence results between Round 1 and Round 2, as should be
expected. The differences are not great, however, particularly when comparing the proportions of
HCBD February 2003
4-12
-------
systems affected. The results range from 0.2-0.4% of PWSs with detections of HCBD and range
from 0.02-0.1% of PWSs with detections greater than the Health Reference Level of 0.9 |ig/L.
These are not substantively different, given the data sources.
The differences in the population extrapolations appear greater, but still constitute relatively
small proportions of the population. The most pronounced difference is in the estimate of the
population served by PWSs with detections greater than the Health Reference Level, ranging from
10,000 to 780,000. In both cases, this is less than 0.5% of the population. The difference in this
category is largely driven by the Florida data in Round 1, as discussed above.
The Round 2 cross-section provides a better proportional balance related to the national
population of PWSs and may have fewer reporting problems than Round 1 (i.e., incomplete
population numbers, Florida). The larger estimate of the national population served by PWSs with
detections greater than the Health Reference Level using Round 1 data can also provide an upper
bound estimate in considering the data.
Regional Patterns
Occurrence results are displayed graphically by State in Figures 4-2, 4-3, and 4-4 to assess
whether any distinct regional patterns of occurrence are present. Combining Round 1 and Round 2
data (Figure 4-2), there are 47 States reporting. Six of those States have no data for HCBD, while
another 21 have no detections of the chemical. The remaining 20 States have detected HCBD in
drinking water and are well distributed throughout the United States.
The simple spatial analysis presented in Figures 4-2, 4-3, and 4-4 suggests that special
regional analyses are not warranted. Florida's possible bias is notable, however. While no clear
geographical patterns of occurrence are apparent, comparisons with environmental use and release
information are useful (see also Section 2.2). Five of the eight Toxic Release Inventory States that
reported releases of HCBD into the environment between 1988 and 1998 have also detected the
chemical in PWS sampling. Of the remaining three (Kansas, Louisiana, and California), Kansas
hasn't reported any data for either Round 1 or 2. Also, of the eight States with detections of HCBD
at CERCLA National Priorities List (NPL) hazardous waste sites, five have detected the chemical
in drinking water. Finally, six of the States detecting HCBD in PWS samples have also detected it
in site samples reported to the ATSDR's HazDat database. It is interesting to note that neither
Alabama nor Florida, the two States with the highest percentage of PWSs with detections greater
than the Health Reference level, are Toxic Release Inventory (TRI) States for HCBD, nor do they
have CERCLA NPL sites with detections of the chemical (Figure 4-4).
HCBD February 2003 4-13
-------
Figure 4-2. States with PWSs with Detections of Hexachlorobutadiene for all States with
Data in URCIS (Round 1) and SDWIS/FED (Round 2).
All States
Hexachlorobutadiene Detections
in Round 1 and Round 2
States not in Round 1 or Round 2
No data for Hexachlorobutadiene
States with No Detections (No PWSs > MRL)
1 States with Detections (Any PWSs > MRL)
HCBD February 2003
4-14
-------
Figure 4-3. States with PWSs with Detections of Hexachlorobutadiene (any PWSs with
results greater than the Minimum Reporting Level [MRL]) for Round 1 (above)
and Round 2 (below) Cross-section States.
* State of Florida is an outlier with 5.36% PWS > MRL
Hexachlorobutadiene Occurrence in Round 1
| | States not in Cross-Section
| | No data for Hexachlorbutadiene
| | 0.00% PWSs > MRL
| | 0.01 - 1.00% PWSs > MRL
r~m i.oo - 3.50% PWSS > MRL*
Hexachlorobutadiene Occurrence
in Round 2
States not in Cross-Section
No data for Hexachlorbutadiene
0.00% PWSs > MRL
0.01 - 1.00% PWSs > MRL
1.00-3.50% PWSs > MRL
HCBD February 2003
4-15
-------
Figure 4-4. Cross-section States (Round 1 and Round 2 combined) with PWSs with
Detections of Hexachlorobutadiene (above) and concentrations greater than the
Health Reference Level (HRL; below).
* State of Florida is an outlier with 5.36%PWS >MRL
Hexachlorobutadiene Occurrence
in Round 1 and Round 2
| | States not in Cross-Section
| | No data for Hexachlorbutadiene
I | 0.00% PWSs > MRL
| | 0.01 - 1.00% PWSs > MRL
r~m i.oo - 3.50% PWSS > MRL*
Hexachlorobutadiene Occurrence
in Round 1 and Round 2
| States not in Cross-Section
| No data for Hexachlorobutadiene
^H 0.00% PWSs > HRL
| 0.01-1.00% PWSs > HRL
"" 1.00 -3.50% PWSs > HRL*
HCBD February 2003
4-16
-------
4.3 Conclusions
While there have not been detections of the chemical in ambient water reported in USGS
NAWQA studies to date, hexachlorobutadiene has been detected at a very low percentage of
ATSDR HazDat sites and CERCLA NPL sites. Furthermore, releases have been reported through
the TRI.
Hexachlorobutadiene has also been detected in PWS samples collected under SDWA.
Occurrence estimates are low for Round 1 and Round 2 monitoring with only 0.13% and 0.05% of
all samples showing detections, respectively. Significantly, the values for the 99th percentile and
median concentrations of all samples are less than the Minimum Reporting Level. For Round 1
samples with detections, the median concentration is 0.25 |ig/L and the 99th percentile concentration
is 10 |ig/L. Median and 99th percentile concentrations for Round 2 detections are 0.30 |ig/L and 1.5
Hg/L, respectively. Systems with detections only constitute 0.4% of Round 1 systems and 0.2% for
Round 2. National estimates for the population served by PWSs with detections are also low,
especially for detections greater than the Health Reference Level. For both rounds, these estimates
are less than 0.5% of the national population (Round 1: 781,076; Round 2: 9,721).
HCBD February 2003 4-17
-------
5.0 EXPOSURE FROM MEDIA OTHER THAN WATER
This section describes studies which measured concentrations of HCBD in food, air, and soil.
Exposure of adults and children is estimated by combining the reported concentrations with the
estimated intake of each medium. These calculations enable a comparison of exposure to HCBD
from air, food, and soil with that anticipated from ingestion of drinking water (see Chapter 9.0).
Estimates of human exposure to HCBD via food and air have previously been calculated by U.S.
EPA(1998a).
5.1 Exposure from Food
Food may be contaminated with HCBD via environmental sources or by contact with
contaminated water during food processing activity (DiNovi, 1997). According to the Food and
Drug Administration (FDA), there are no approved uses of HCBD either directly or indirectly in
foods, including food processing equipment (DiNovi, 1997). HCBD is not regulated in plastics.
5.1.1 Concentrations in Non-Fish Food Items
Two reports provide data for the concentration of HCBD in food items. Yip (1976)
measured HCBD in food items within a 25-mile radius of tetrachloroethylene and trichloroethylene
manufacturing plants that emit HCBD as a waste product. No HCBD was detected in 15 egg samples
and 20 vegetable samples. One of 20 milk samples contained 1.32 mg/kg HCBD. Resampling in the
same area revealed no further detections in milk, raising the possibility that the concentration of 1.32
mg/kg measured in the original data set was an artifact. This study reported two detection limits for
HCBD: 0.005 mg/kg for nonfatty foods and 0.04 mg/kg for fatty foods. Based on information
supplied by Kusznesof (1997), U.S. EPA (1998a) concluded that more than 30% of foods may be
considered fatty foods for the purpose of estimating exposure from food (see Section 5.1.3).
IARC (1979) reported concentrations of HCBD in foods sampled in the United Kingdom.
HCBD was found at concentrations of 0.00008 mg/kg in fresh milk, 0.002 mg/kg in butter, 0.0002
mg/kg in cooking oil, 0.0002 mg/kg in light ale, 0.0008 mg/kg in tomatoes, and 0.0037 mg/kg in
black grapes (IARC, 1979).
5.1.2 Concentrations in Fish
Concentrations of HCBD in fish have been reported in multiple studies. Tchounwou et al.
(1998) demonstrated that aquatic organisms, particularly fish, may be a significant source of HCBD
transmission from contaminated wetlands to humans. Tissue concentrations of HCBD were 226.33
± 778.40 ng/g in fish collected from a contaminated study site in Louisiana and 6.84 ± 10.41 ng/g
in fish collected from the corresponding control site.
In other studies, fish samples from the Mississippi River were reported to contain HCBD
levels ranging from 100 to 4,700 ng/g (Laska et al., 1976; Yip, 1976; Yurawecz et al, 1976). Levels
of HCBD generally were not detected in fish from the Great Lakes (Camanzo et al., 1987; DeVault,
1985), with the exception of trout from Lake Ontario, which were reported to contain 60 to 300 ng/g
HCBD February 2003 5-1
-------
(Oliver and Nimi, 1983). HCBD was not detected in 51 biota samples catalogued in the STORET
database (Staples et al., 1985).
The National Study of Chemical Residues in Fish (NSCRF), conducted by EPA's Office of
Water, was undertaken to determine the occurrence of selected pollutants in fish from various
locations across the United States. Pollutants were measured in bottom-feeding and game fish at
nearly 400 sites between 1986 and 1989 (Kuehl et al., 1994). A complete presentation of the study
plan and results is contained in a joint Office of Water and Office of Research and Development
report (U.S. EPA, 1992a). To obtain nationwide coverage, samples were collected at sites near
potential point and nonpoint pollution sources, at background sites in areas generally without
pollution sources, and at a few sites from the U.S. Geological Survey's National Stream Quality
Accounting Network (NASQAN). Targeted sites were chosen near areas of significant industrial,
urban, or agricultural activities, including more than 100 sites near pulp and paper mills.
Fish species chosen for sampling were those routinely consumed by humans and/or those
expected to bioaccumulate organic contaminants. At most locations, the NSCRF analyzed one
composite sample of bottom-feeding fish, usually composed of whole-body samples. Some bottom-
feeding fish composite samples were composed of fillets. In areas where whole-body concentrations
were high, composite samples of game fish were usually composed of fillets. Each composite sample
contained approximately three to five adult fish of similar size from the site. Pollutant concentrations
were measured in units of wet weight (U.S. EPA, 1992a).
HCBD was detected in fish at 3% of the 362 sites sampled. Fillet samples were taken from
106 sites. The mean and standard deviation of HCBD fish concentrations at all sites were 0.6 ng/g
and 8.7 ng/g, respectively (Kuehl et al., 1994). These statistics represent the overall mean from all
samples, not just from the positive samples. Concentrations were above 2.5 ng/g at only four sites,
which were all near organic chemical manufacturing plants (U.S. EPA, 1992a). The concentrations
observed at these four sites are provided in Table 5-1.
The methods for determining the mean and standard deviation for HCBD concentration and
for evaluating samples below the analytical detection limit were not specifically stated by U. S. EPA
(1992a). The value of the detection limit for HCBD was not given in U.S. EPA (1992a) or Kuehl
et al. (1994). However, in the Kuehl et al. (1994) study, the mean concentration was calculated using
one-half of the detection limit concentration when the analyte was not detected. The raw data for
HCBD were not presented.
Hendricks et al. (1998) evaluated HCBD levels in zebra mussel (Dreissenapolymorphd) and
eel {Anguilla anguilla) from approximately 30 locations in the Rhine-Meuse river basin. In zebra
mussel, HCBD levels were 240 ng/kg at a background location and ranged from 950 to 14,000 ng/kg
wet weight within the study area. In eel, HCBD levels were found to range from 5,000 to 55,000
ng/kg wet weight within the study area.
HCBD February 2003 5-2
-------
Table 5-1. HCBD Tissue Concentration in Fish Collected Near Four Chemical
Manufacturing Plants.
CONCENTRATION
(ng/g wet weight)
164.0
23.0
10.50
2.54
TYPE OF SAMPLE
Sea Catfish - Whole Body
Sea Catfish - Whole Body
Catfish - Fillet
Catfish - Whole Body
LOCATION
Louisiana
Texas
Illinois
Louisiana
source: U.S. EPA (1992a)
5.1.3 Intake of HCBD from Food
Non-fish Dietary Intake
As noted above, HCBD has been found in a variety of foods in the United Kingdom. In
addition, HCBD may have been incorrectly measured in one milk sample in the study by Yip (1976).
It is also possible that HCBD could be found in measurable quantities in the United States.
However, because HCBD was generally undetected in samples taken from areas within 25 miles of
emission sources, U.S. EPA (1998a) concluded that it is appropriate to assume that, on average,
HCBD will not be found in food at detectable levels. Given this observation, along with the fact that
HCBD has no approved uses in food, it is anticipated that there would typically be no chronic
exposure to HCBD via non-fish dietary foods (U.S. EPA, 1998a). Therefore, the average estimate
of HCBD intake from non-fish foods is assumed to be zero (U.S. EPA, 1998a).
A high-end estimate of HCBD exposure may be made by assuming a concentration of one-
half the detection limit (U.S. EPA, 1999b). Because the percentages of fatty or non-fatty foods in
the diet are not known with certainty, a conservative estimate is made using one-half the detection
limit of 0.04 mg/kg noted for fatty foods in Yip (1976). The resulting concentration of 0.02 mg/kg
is multiplied by an estimate of total daily food intake of 2.6 kg/day and divided by 70 kg to obtain
a total daily intake of HCBD from food of 7.4 x 10"4 mg/kg-day in adults. For children, the resulting
concentration of 0.02 mg/kg was multiplied by an estimate of total daily food intake of 0.84 kg/day
(U.S. EPA, 1988) and divided by a body weight of 10 kg to obtain a total daily intake of HCBD
from food of 1.68 x 10"3 mg/kg-day. For the maj ority of regions of the United States in which HCBD
is not found, using one-half the detection limit will overestimate the amount of HCBD in food (U. S.
EPA, 1999b).
Because the data on concentrations in food are limited, and because the implications of
assuming that HCBD occurs at one half the detection limit for fatty foods are large, further research
may be required to refine this estimate.
HCBD February 2003
5-3
-------
Fish Dietary Intake
U. S. EPA (1998a) estimated HCBD intake from fish using the tissue concentration data from
Kuehl et al. (1994). Because these data were taken from many monitoring stations throughout the
United States, the estimate may be reasonably indicative of the magnitude of intake from fish
consumption when HCBD is present in fish tissue. An average estimate of adult exposure was
obtained by multiplying the mean concentration of 0.6 ng/g from the Kuehl et al. (1994) data by a
fish intake of 18 g/day for the general population and dividing by a body weight of 70 kg. The
resulting estimate is 1.54 x 10"7 mg/kg-day. The maximum concentration detected in fish by Kuehl
et al. (1994) can also be used to estimate the high-end intake. Following the same procedure above,
but substituting a concentration of 164 ng/g, one obtains a high-end intake of 4.22 x 10"5 mg/kg-day.
An average estimate of HCBD exposure in children was determined by multiplying the mean
concentration of 0.6 ng/g from the Kuehl et al. (1994) data by a fish intake of 4 g/day for the general
population and dividing by a body weight of 10 kg. The resulting estimate is 2.4 x 10"7 mg/kg-
day. The maximum concentration detected in fish by Kuehl et al. can also be used to calculate a
high-end estimate of intake in children. Following the same procedure above, but substituting a
concentration of 164 ng/g, results in an intake of 4.37 x 10"5 mg/kg-day.
5.2 Exposure from Air
5.2.1 Concentration of HCBD in Air
Concentration data for HCBD in air have previously been summarized by U. S. EPA (1998a).
The largest compilation of data on ambient air concentrations is available from Shah and Heyerdahl
(1988). Shah and Heyerdahl compiled ambient air monitoring data for volatile organic compounds
for the period from 1970 to 1987. A total of 72 observations from six studies were reported for
HCBD. In cases where more than one sample was taken per day, the concentrations were averaged
and weighted by sampling time when the sampling time varied throughout the day. When more than
one sample was included in the average, values less than the minimum quantifiable limit (MQL)
were included as one-half the MQL when the MQL was given. When the MQL was not indicated
in the Shah and Heyerdahl study, values less than the MQL were included as zeros in the average.
If the resulting average was less than the MQL, a zero was included. If the average was greater than
the MQL, the calculated average was used.
As reported in U.S. EPA (1998a), the average and median of all ambient HCBD air
concentrations measured by Shah and Heyerdahl (1988) were 0.036 parts per billion (ppb) (0.42
l-ig/m3) and 0.003 ppb (0.04 |ig/m3), respectively. The 25th and 75th percentiles were 0.001 ppb
(0.01 jig/m3) and 0.006 ppb (0.07 |ig/m3). Only median values were reported for urban areas and
source-dominated areas. Of 56 samples taken from urban areas, the median was 0.003 ppb (0.04
l-ig/m3). Of 16 samples taken from source-dominated areas, the median was 0.002 ppb (0.02 |ig/m3).
No indoor concentrations were reported (Shah and Heyerdahl, 1988).
Shah and Heyerdahl's compilation included the results from Pellizzari et al. (1979), who
surveyed the occurrence of halogenated hydrocarbons in various environmental media of five
metropolitan areas. As part of this study, HCBD concentrations in the vapor phase of ambient air
HCBD February 2003 5-4
-------
of four sites were compiled from other research programs, as well as from monitoring conducted
specifically for this project. In the Niagara Falls and Buffalo, New York area, concentrations were
found to range from trace levels to 389 ng/m3, with six of 15 determinations (40%) containing
detectable levels. In the Baton Rouge, Louisiana area, two of 11 determinations (18%) were 18 and
37 ng/m3. Sampling in Houston, Texas, and surrounding areas showed a range of trace levels to
2,066 ng/m3, with seven positive values from a total of 17 determinations (41%).
Class and Ballschmiter (1987) reported that the troposphere of the Northern Hemisphere
contained an average concentration of 0.17 parts per trillion (ppt) (2 jig/m3) HCBD at 18 locations
sampled from 1982 to 1986. The detection limits in this survey were between 0.01 and 0.1 ppt.
HCBD concentrations in ambient air were measured in two studies included in a compilation
of ambient monitoring data for the Urban Area Source Program (U.S. EPA, 1994). In the first
survey, concentrations of HCBD were reported at a minimum detection level of 540 |ig/m3 when
measured at six monitoring stations in Columbus, Ohio, in 1989. The second survey was conducted
in Cincinnati, Ohio, from 1989 to 1991, and detected HCBD at one monitoring site at a
concentration of 1,000 |ig/m3.
A number of cities had HCBD levels ranging from 2 to 11 ppt (0.02 to 0.12 |ig/m3)
(Pellizzari, 1978; Singh et al., 1980, 1982). Niagara Falls had higher HCBD levels, with
concentrations up to 37 ppt (0.39 |ig/m3) found in ambient air levels and up to 38 ppt (0.41 |ig/m3)
found in the basement air of homes near industrial and chemical waste disposal sites (Pellizzari,
1982).
However, a study of air contaminants in Porto Alegre, Brazil (Grosjean and Rassmussen,
1999) did not find detectable levels of HCBD (detection limit = 100 ppt) at any of 46 sampling
locations. A monitoring study at 6 sampling locations in Columbus, Ohio also failed to detect HCBD
in the air (Spicer et al., 1996).
5.2.2 Intake of HCBD from Air
The air concentrations reported in Shah and Heyerdahl (1988) were utilized by U.S. EPA
(1998a) to calculate an estimate of exposure because this data set included a reasonable number of
observations (n=72). For adults, the mean concentration of 0.42 |ig/m3 was multiplied by an average
air intake of 20 m3/day (U.S. EPA, 1988). The resulting value was divided by a body weight of 70
kg, and the units were converted from |ig to mg, resulting in an average intake of 1.2 x 10"4 mg/kg-
day. For children, the mean concentration of 0.42 |ig/m3 was multiplied by an average air intake of
15 nrVday (U.S. EPA, 1988). The resulting value was divided by a body weight of 10 kg and the
units were converted from |ig to mg, resulting in an intake of 6.3 x 10"4 mg/kg-day. As noted in U. S.
EPA (1998a), these estimates may be indicative of the magnitude of HCBD intake from air in urban
and source dominated areas where the chemical is present. It should be noted, however, that these
concentration data are older than data from the Urban Area Source Program (U.S. EPA, 1994) and
Class and Ballschmiter (1987). In addition, the number of geographic areas sampled throughout the
United States by Shah and Heyerdahl was not indicated.
HCBD February 2003 5-5
-------
5.3 Exposure from Soil
5.3.1 Concentration of HCBD in Soil and Sediment
Limited data on the concentrations of HCBD in soil was identified in a RCRA corrective
measures study, and ranged from 0.043 - 0.35 mg/kg (USDOE, 2001) . Sediments adsorb HCBD
from contaminated water. As reported in U.S. EPA (1999b), HCBD was not detectable in any of the
196 sediment samples reported in the STORET database, based on a detection limit of 500 jig/kg
for the analyses (Staples et al., 1985). Sediments from the Niagara River contained 2.9 to 11 i-ig/kg
HCBD (Oliver and Bourbonniere, 1985). Sediments from the Great Lakes contain HCBD typically
ranging from 0.08 to 120 |ig/kg (McConnell et al., 1975). Suspended solids collected from the
Rhine-Meuse river basin indicate HCBD levels ranging from < 3.4 to 19 |ig/kg (Hendriks et al.,
1998); which is comparable to the previously mentioned data for sediments collected in the United
States.
Several studies have investigated HCBD levels in sediments from sites in Louisiana. HCBD
levels in sediment samples from a Louisiana swamp environment ranged from less than 0.05 |ig/kg
to 0.40 i-ig/kg (Abdelghani et al,. 1995). These concentrations were well below the action levels of
4,000 jig/kg for sediment (U.S. EPA, 199la). At a Federal Superfund site near Baton Rouge,
Louisiana, preliminary data from a sampling of sediments showed HCBD levels from 2 to 3,770
mg/kg (U. S. EPA, 1992b). The HCBD level in a sediment sample from Lake Charles, Louisiana was
found to be 3,500 jig/kg (Chen et al., 1999). A sediment sample collected from the intersection of
an industrial canal and Bayou d'Inde (a tributary of the Calcasieu River near Lake Charles) and
analyzed via Soxhlet extraction was found to contain HCBD at a level of 17,200 ± 1,000 jig/kg
(Prytula and Pavlostathis, 1996). Another reported sediment sample collected from this industrial
canal area had an HCBD level of 36,000 ± 6,900 jig/kg (Gess and Pavlostathis, 1997). A third study
of sediments from Bayou d'Inde found levels of HCBD ranging from 1,550 to 8,220,000 |ig/kg of
organic carbon. Assuming an organic carbon content in the sediment of 1%, this level is equivalent
to sediment concentrations of 15 to 82,200 jig/kg.
5.3.2 Intake of HCBD from Soil
Because no data were available on the concentration of HCBD in soil or dust, intake from
soil was not estimated.
5.4 Other Residential Exposures
HCBD was not detected in sewage influents (Levins et al., 1979) or in sewage samples (U. S.
EPA, 1990). No other information on exposure via other potentially complete residential pathways
was identified.
5.5 Summary
Estimated mean concentration and average intake values for HCBD in media other than
water are summarized in Table 5-2. Assuming that there is no chronic exposure of the general
population
HCBD February 2003 5-6
-------
Table 5-2. Summary of Concentration Data and Exposure Estimates for Media Other
Than Water.
PARAMETER
Mean Concentration in medium
Estimated average daily intake (mg/kg-day)
MEDIUM
Food
Non-fish (NF): nondetect
Fish(F): 0.6ng/g
Adult
NF:0
F: 1.5 x 10'7
Child
NF:0
F: 2.4 x 10'7
Air
0.42 ng/m3
Adult
1.2 x 10'4
Child
6.3 x lO'4
to HCBD from non-fish dietary sources, inspection of the data indicates that most intake of HCBD
by the general population occurs via inhalation. However, it should be cautioned that this
preliminary conclusion is subject to a number of uncertainties: 1) the database for the occurrence
of HCBD in media other than water is limited; 2) many of these data are more than 20 years old; 3)
in some cases, information on the geographic location of sample collection or analytical details are
lacking; and 4) data for HCBD in soil and dust were not available to estimate via this pathway.
HCBD February 2003
5-7
-------
6.0 TOXICOKINETICS
This section describes the absorption, distribution, metabolism, and excretion of
hexachlorobutadiene. The information in this section focuses on findings in animals exposed
primarily via the oral route. No studies were identified that evaluated the toxicokinetic behavior of
HCBD in humans. The development of a physiologically based pharmacokinetic (PBPK) model
would be desirable for summarizing and extrapolating the toxicokinetic behavior of HCBD between
high-dose animal studies and low-dose human exposure. In order to develop such a model, blood
or plasma time-course data for HCBD and, if possible, its metabolites, would be required. However,
no pharmacokinetic parameters (half-life, partition coefficients, etc.) for HCBD or its metabolites
have been published in the peer-reviewed literature, and no blood or plasma time-course for HCBD
or its metabolites has been published for any species. This remains as an important gap in the
database for this chemical.
6.1 Absorption
HCBD is readily absorbed following oral administration to experimental animals. Although
no studies have quantitatively determined the rate of absorption of HCBD following oral dosing,
useful information has been obtained from studies that evaluated the distribution and excretion of
this compound. Reichertetal. (1985) administered 1 mg/kgof 14C-HCBD to female Wistar rats via
gavage. The compound was administered in a tricaprylin suspension to accommodate its low water
solubility. Approximately 76% of the radioactivity was excreted as metabolites in the urine, feces,
or expired air within 72 hours after administration, suggesting that most of the dose was absorbed.
When a higher dose of 50 mg/kg 14C-HCBD was administered in the same study, 69% of the
radioactivity was found in the feces and was predominantly associated with unchanged HCBD. Just
11% of the administered radioactivity was excreted in the urine for the high-dose group, compared
to 31% for the low-dose animals. The study authors concluded that absorption of HCBD was
saturated in animals in the higher-dose group (Reichert et al. 1985; U.S. EPA, 1991a).
Nash et al. (1984) administered 200 mg/kg 14C-HCBD via oral gavage in corn oil to male
Wistar-derived rats. Animals were sacrificed 2, 4, 8, or 16 hours after dosing, and the fate of the
administered radioactivity was evaluated using whole-body autoradiographs. The investigators
reported that absorption was virtually complete within 16 hours after dosing.
Payan et al. (1991) administered 1 mg/kg and 100 mg/kg 14C-HCBD to male Sprague-
Dawley rats, using an aqueous polyethylene glycol vehicle, and found that 18.5 and 8.9% of the
administered radioactivity, respectively, was excreted over 72 hours in the urine. Since urinary
excretion at a dose of 1 mg/kg in the Reichert et al. (1985) study was 31%, these data suggest that
gastrointestinal absorption of HCBD was greater when administered in a lipophilic vehicle
(tricaprylin) than with an aqueous vehicle (aqueous polyethylene glycol), although it could also be
due to differences in animal sex and strain (female Wistar rats vs. male Sprague-Dawley rats). As
noted for other unsaturated chlorinated compounds, HCBD absorption presumably occurs by passive
diffusion across the lipid portion of the intestinal matrix rather than by active or protein-facilitated
transport (ATSDR, 1994).
Little information is available regarding HCBD absorption following exposure by other
routes. Although no studies were located that described absorption in humans or animals after
HCBD February 2003 6-1
-------
inhalation exposure to this compound, the occurrence of systemic effects following exposure
indicates that absorption occurs by this route (ATSDR, 1994). With regard to dermal exposure,
Duprat and Gradiski (1978) applied doses of 419 to 1,675 mg/kg HCBD to the skin of rabbits under
occluded conditions, and reported that the compound was completely absorbed within 8 hours.
6.2 Distribution
HCBD has been detected in the adipose tissue of human cadavers at concentrations of 0.003
± 0.001 |ig/g (Mes et al., 1985), but not whole blood in any of 36 residents of Love Canal, New
York or in any of 12 laboratory volunteers (Bristol et al., 1982). Olea et al. (1999) detected HCBD
in the adipose tissue of 13 of 50 children living in an agricultural region of southern Spain. The
mean concentration in the 13 children was 0.70 |ig HCBD/g of fat (range: 0.23 to 2.43 |ig HCBD/g
of fat). No data were available concerning the route of exposure.
Following oral administration, HCBD-related radioactivity preferentially distributed to the
kidney, liver, adipose tissue and brain of experimental animals (Reichert, 1983; Rei chert etal., 1985;
Dekant et al., 1988a). Covalent binding of HCBD-related radioactivity to tissue proteins in female
Wistar rats was highest during the first six hours after dosing, and was higher in the kidney than in
the liver. This effect was independent of dose (Reichert et al., 1985). In female Wistar rats
administered 1 mg/kg 14C-HCBD, covalent binding of the radioactivity to protein in the kidney was
about twice that in the liver 72 hours after dosing (Reichert et al., 1985). Nash et al. (1984) reported
a specific localization of administered radioactivity in male Wi star-derived rats in the outer medulla
of the kidney, as revealed by autoradiographic analysis following an oral dose of 200 mg/kg 14C-
HCBD. Payan et al. (1991) conducted a study in male Sprague-Dawley rats in which the bile ducts
of one group of animals administered an oral dose of 100 mg/kg 14C-HCBD were cannulated so that
bile secretions from these animals could be infused directly into the duodenum of another group of
animals. In both groups, the kidneys contained about twice as much radiolabel as the liver.
No studies were located regarding the distribution of HCBD in humans or animals after
inhalation or dermal exposure. Davis et al. (1980) administered 0.1 mg/kg radiolabeled 14C-HCBD
as a tracer dose to a control group of male Sprague-Dawley rats (5 animals/group) via intraperitoneal
injection. Another group received the same amount of labeled HCBD plus a nephrotoxic dose of
300 mg/kg non-labeled HCBD. The highest concentrations of radiolabel were found in the liver,
kidney, and adipose tissue 48 hours after administration. Approximately 2.6 and 2.3% of the
administered 14C radiolabel were retained in the livers of low- and high-dose animals, respectively.
The fraction of the tracer retained in kidney varied from 2.5% at the low dose to 0.5% at the high
dose. The fraction of the dose found in adipose tissue was not determined. Very low levels of the
radiolabel (less than 0.2%) were found in the brain, lung, heart, and muscle.
6.3 Metabolism
No available studies have characterized the metabolism of HCBD in animals following
inhalation or dermal exposure. The metabolism of HCBD in animals has been studied in isolated
hepatocytes from male Sprague-Dawley rats (Jones et al., 1985) and by characterization of
metabolites identified in urine, bile, and feces following oral exposure to the compound (Figure 6-1).
Following ingestion and absorption from the gastrointestinal tract of mice and rats, HCBD is initially
HCBD February 2003 6-2
-------
transported to the liver, where it is conjugated with glutathione to form £-(1,1,2,3,4 -
pentachlorobutadienyl)glutathione in a reaction mediated by glutathione S-transferase (Wolf et al.,
1984; Garle and Fry, 1989; Dekant et al., 1988b; Koob and Dekant, 1992). No studies have found
non-enzymatic glutathione-conjugation of HCBD, and none have investigated GST subtype
specificity for HCBD. In male Sprague-Dawley rats, a di-substituted glutathione conjugate, 1,4
bis(l,2,3,4-tetrachlorobutadienyl)glutathione, is also formed in the liver (Jones et al., 1985),
whereas in mice, only the mono-substituted conjugate is produced (Dekant et al., 1988a). The
glutathione conjugate is then excreted in bile and transported back into the gastrointestinal tract
(Koob andDekant, 1992). Nash etal. (1984), for example, collected bile excretions from cannulated
male Wistar-derived rats that had been orally administered 200 mg/kg 14C-HCBD and determined
that 40% of the bile radioactivity was associated with the glutathione conjugate. Studies conducted
by Lock et al. (1984) demonstrate that induction or inhibition of cytochrome P450 metabolism does
not alter the nephrotoxicity of HCBD in male or female mice administered 24, 48, 96, or 144
mol/kg HCBD by gavage. Glutathione and N-acetylcysteine conjugates of HCBD exhibit more
potent nephrotoxicity in male and female mice (Lock et al., 1984; Ishmael and Lock, 1986) and male
Sprague-Dawley rats (Nash et al., 1984) than unconjugated HCBD. This suggests that conjugated
metabolites, unlike oxidative metabolites, may play a large role in HCBD nephrotoxicity.
The glutathione conjugate of HCBD can be reabsorbed intact from the gastrointestinal tract
of male rats (Koob and Dekant, 1992; Gietl et al., 1991). Alternatively, a portion of it can be
catabolized by y-glutamyltranspeptidase and dipeptidases in the rat gastrointestinal tract to the
cysteine conjugate, ^-(IJ^S^-pentachlorobutadieny^-L-cysteine (Jones etal., 1985; Gietl etal.,
1991; Koob and Dekant, 1992). In the rat, both the glutathione and cysteine conjugates are subject
to several alternative fates. These conjugates may be reabsorbed from the gut and be translocated
to the kidney (Koob and Dekant, 1992), undergo enterohepatic circulation (Nash et al., 1984; Gietl
etal., 1991; Gietl and Anders, 1991), or be excreted with the feces (Dekant etal. 198 8a). However,
the majority of the glutathione conjugate is delivered to the kidney by systemic circulation in the rat
(Koob and Dekant, 1992). Working with isolated perfused rat livers, Koob and Dekant (1992)
determined that a maximum of 39% of the glutathione conjugate was recirculated to the liver of rats,
whereas up to 79% of the cysteine conjugate was recirculated. Nash et al. (1984) reported that the
cysteine and glutathione conjugates represented 12% and 40%, respectively, of the radioactivity
excreted in the bile of cannulated rats orally administered 200 mg/kg HCBD. When the cysteine
conjugate is recirculated to the rat liver, a minor fraction of this metabolite is converted by N-
acetyltransferase to an acetylated cysteine conjugate, N-acety\-S-( 1,1,2,3,4-pentachlorobutadienyl)-
L-cysteine (7V-AcPCBC) (Koob and Dekant, 1992).
HCBD February 2003 6-3
-------
Figure 6-1. Proposed Pathways for Hexachlorobutadiene Metabolism.
Cl^ /Cl Ck xCI
Ck yCl
r 1 V :r- r i \ MH^^^H Cyt' P45° 3A ^ f \ ' Mlirnril
/A / \ 1 Male-specific ^ \© 1
Cl Cl Cl XS-CH2-CH Sulfoxidation pathway r( ^S-CH?-CH
Hexachloro-1,3-butadiene N-acetyl-S-(1,1,2,3,4-pentachloro- 1 1 COOH
Glutathione fiver^./ butadienyl)-L-cysteine (N-AcPCBC)
Transferase y N-acetyl-transferase 1 deacetylase
Cl /C\ * -glutamic acid \ /* _Qlycine ck ^
\ / 1-alutamvl -C C Cl ^- \ /
/> \ /* transpeptidase cf V=C^ Cl^~^\ /°' NH' '
/ C\G C' S^Cys^Gly / \^^
S-(1,1,2,3,4-pentachloro I
butadienyl)glutathione (PCBG) COOH
& I
©O COOH
N-acetyl-S-(1,1,2,3,4-pentachlorobutadienyl)
-L-cysteine sulfoxide (N-AcPCBC-SO)
deamination^ ck /cl
iecarboxylation ,-* Q C|
Cl /C==C\
Cl S-CH2-COOH
1 ,1 ,2,3,4-pentachlorobutadienyl
,, / S-(1,1,2,3,4-pentachlorobutadienyl)- carboxymethylthioether
Glutathione
-------
Further processing of both the cysteine and glutathione conjugates occurs in the kidney,
which possesses high y-glutamyltranspeptidase activities in the brush-border membrane of the
proximal tubular cells (Dekant and Vamvakas, 1993; Dekant et al., 1990). Renal deacetylase, y-
glutamyltranspeptidase, and dipeptidase enzymes convert the acetylated cysteine conjugate and the
glutathione conjugate to the cysteine conjugate, which accumulates in the kidney (Dekant et al.,
1990). The cysteine conjugate is subsequently activated to a reactive and electrophilic thioketene
intermediate (Dekant et al., 1990; Green and Odum 1985). This conversion is catalyzed by the
enzyme-cysteine conjugate P-lyase, which is localized in the cytosol and mitochondria of the
epithelial cells of the proximal tubule (Lash et al., 1986; Stevens, 1985; Stevens et al. 1986; Jones
et al., 1988; MacFarlane et al., 1989; Kim et al., 1997).
Another pathway for metabolic disposition of the cysteine conjugate of HCBD in the kidney
is the conversion of the cysteine conjugate to a mercapturic acid, TV-acetyl-^l, 1,2,3,4-
pentachlorobutadienyl)-L-cysteine, by the renal enzyme 7V-acetyltransferase (Birner et al., 1997).
This metabolite is excreted in the urine, accounting for 10% of urinary radioactivity in rats orally
administered 100 mg/kg 14C-HCBD (Reichert and Schutz, 1986). Other pathways that result in the
excretion of the cysteine conjugate involve the deamination and subsequent decarboxylation of the
cysteine conjugate, resulting in the formation of methylthiolated metabolites such as 1,1,2,3,4-
pentachlorobutadienemethylthioetherand 1,1,2,3,4-pentachlorobutadienecarboxymethylthioether
(Reichert et al., 1985). 1 to 8% of the administered radioactivity is recovered as carbon dioxide in
exhaled air from rats (Reichert et al., 1985; Payan et al., 1991), probably from decarboxylation of
the cysteine conjugate.
Evidence for a male-specific HCBD metabolic pathway in rats has been reported by Birner
and colleagues (Birner etal., 1995,1998; Werner etal.!995a). The metaboliteTV-acetyl-^l, 1,2,3,4-
pentachlorobutadienyl)-L-cysteine sulfoxide (7V-AcPCBC-SO) is detected in the urine of male, but
not female, rats following oral administration of HCBD. Formation of this metabolite is mediated
by cytochrome P450 3A monooxygenases, which are expressed only in male rats (Birner et al.,
1995; Werner et al., 1995a). This metabolite has been found to be cytotoxic to proximal tubular cells
in vitro without activation by P-lyase (Birner et al., 1995). When given intravenously, 7V-AcPCBC-
SO produced necroses of the kidney tubules in male rats (Birner et al., 1998).
The TV-AcPCBC-SO formed in male rats occurs as two diastereomers present in equimolar
amounts: (^)-TV-AcPCBC-SO and (S)-N-AcPCEC-SO (Werner etal., 1995b). These compounds are
structurally analogous to unsaturated carbonyl compounds and thus may be candidates for
detoxification via glutathione conjugation (Rosner et al., 1998). Experimental evidence obtained in
vitro suggests that glutathione conjugation of the two diastereomers is catalyzed by different
glutathione ^-transferases, resulting in the formation of different products (Rosner et al., 1998).
Incubation of the (^)-sulfoxide diastereomer with rat liver cytosol resulted in formation of(R)-N-
acetyl-,S'-(4-glutathion-)S'-yl-l,2,3,4-tetrachlorobutadienyl)-L-cysteine sulfoxide. Incubation of the
(,S)-sulfoxide produced two glutathione conjugates identified as (^-jV-acetyl-S-^-glutathion-S-yl-
l,2,3,4-tetrachlorobutadienyl)-L-cysteine sulfoxide and (-S)-jV-acetyl-,S'-(2-glutathion-S-yl-1,3,4,4-
tetrachlorobutadienyl)-L-cysteine sulfoxide. In the presence of rat kidney cytosol, only the (S)-N-
acetyl-
-------
|i-class) and kidney (cc-class). This hypothesis was confirmed by product analysis following
incubation of 7V-AcPCBC-SO with purified rat a- and [i-class glutathione S-transferases.
Very little information is available on the toxicokinetic behavior of HCBD in humans.
However, the key steps in the metabolism of HCBD have been examined in vitro using human
tissues. The human liver microsomal glutathione transferase responsible for HCBD conjugation has
been isolated and purified (McLellan et al., 1989), and the microsomal enzyme activity is 40-fold
higher than the activity detected in the cytosol (Oesch and Wolf, 1989). The rate of enzymatic
formation of S-(l,2,3,4,4-pentachlorobutadienyl)glutathione (PCBG) from HCBD in human liver
cytosol is in the same order of magnitude as the rates observed in rat and mouse cytosol (Dekant
etal., 1998).
The enzyme y-glutamyl transpeptidase, which catalyzes the conversion of glutathione S-
conjugates to the corresponding cysteine conjugates, has been detected in human tissues (Shaw et
al., 1978). The kidney-to-liver activity ratio for y-glutamyl transpeptidase in human tissues is
approximately 22, which is comparable to the ratios observed in pig and guinea pig. However, this
ratio is much lower than in rat, where a kidney-to-liver ratio of 875 has been observed (Hinchman
and Ballatori, 1990).
Cysteine conjugate P-lyase has been isolated and purified from human kidney cytosol (Lash
et al., 1990), and the human P-lyase gene has been cloned and expressed (Perry et al., 1995). P-lyase
activity has been demonstrated in the human kidney (Green et al., 1990) and human proximal tubular
cells (Chen et al., 1990). Collectively, these human studies suggest that humans have the ability to
metabolize HCBD to toxic metabolites. However, the activity of HCBD metabolizing enzymes,
particularly renal P-lyase, may be many-fold lower in humans than the corresponding enzymes in
rat (Lock, 1994; Lash et al., 1990; Anders and Dekant, 1998).
Werner et al. (1995b) demonstrated that human liver microsomes are capable of oxidizing
TV-acetyl-^MAS^-pentachlorobutadieny^-L-cysteine to the corresponding sulfoxide (N-
AcPCBC-SO). In contrast to the male-specific formation of 7V-AcPCBC-SO in rats described above,
formation of the sulfoxide was detected in human microsomes prepared from both male and female
donors. Inhibitor studies suggest that formation of the sulfoxide is catalyzed by members of the
cytochrome P450 3A subfamily. Since this subfamily constitutes a major fraction of cytochrome
P450 content in human liver, the formation of the sulfoxide is expected to occur in humans exposed
to HCBD. Incubation of 7V-AcPCBC-SO with purified human glutathione ^-transferase Ml-1 (la-
class) catalyzes the formation of (
-------
up to 100 mg/kg in rats and mice (Reichert and Schutz, 1986; Dekant et al., 1988a). At higher doses
(30 to 200 mg/kg), 5 to 11% of the radiolabel was excreted in the urine (Dekant et al., 1988a; Nash
et al., 1984; Reichert and Schutz, 1986; Reichert et al., 1985; Payan et al., 1991), while a dose of
1 mg/kg resulted in urinary excretion of approximately 18.5 to 30% of the administered radioactivity
(Payan et al., 1991; Reichert et al., 1985). Payan et al. (1991) attributed the fractional decrease in
urinary excretion with increase in dosage to a saturation of hepatobiliary excretion or a reduction
of biliary metabolite reabsorption. Exhaled unchanged labeled HCBD accounted for approximately
5% of oral doses of 1 and 50 mg/kg when measured within 72 hours of administration, while exhaled
labeled carbon dioxide accounted for 1% or 4% of the administered dose (Reichert et al., 1985). The
authors did not indicate the metabolic pathway for carbon dioxide formation.
Fecal excretion is the main pathway of elimination for HCBD and HCBD metabolites.
Reichert et al. (1985) reported that elimination of HCBD in the feces represented 42 or 69% of the
radioactivity orally administered to female Wistar rats at doses of 1 or 50 mg/kg, respectively. The
difference in recovery was attributed to an apparent absorption saturation threshold. Dekant et al.
(1988a) found that 67 to 77% of the radioactivity in an oral 30 mg/kg dose in corn oil was present
in the feces of mice 72 hours after administration. Payan et al. (1991) administered oral doses of
14C-HCBD in polyethylene glycol to male rats. After 72 hours, the feces and contents of the
gastrointestinal tract contained 62% of a 1 mg/kg dose and 72% of a 100 mg/kg dose.
Enterohepatic circulation has been demonstrated in animals following oral administration
of HCBD. Nash et al. (1984) administered 200 mg/kg 14C-HCBD to rats with and without cannulated
bile ducts. Feces and urine were collected over a 5-day period. Over the course of the experiment,
the non-cannulated animals excreted 39% of the administered radioactivity in feces. On each of the
first two days post-dosing, approximately 5% and 3.5% of the administered radioactivity were found
in the feces and urine, respectively. In contrast, bile excretions collected from the cannulated
animals on each of the first 2 days post-dosing contained 17-20% of the administered radioactivity
(with 35% total excretion by this route over two days). These findings indicate extensive
reabsorption of biliary metabolites.
Payan et al. (1991) also compared excretion patterns in bile-duct cannulated and
noncannulated rats orally dosed with 1 mg/kg 14C-HCBD. Urinary excretion after 72 hours
accounted for 18% of the administered radioactivity in intact animals, but just 11% of the
radioactivity in the cannulated rats. In comparison, fecal excretion represented 62% and 3%, of the
dose administered to non-cannulated and cannulated animals, respectively. In cannulated rats, 67%
of the dose was excreted into the bile. When bile excretions (isolated from bile duct-cannulated rats
orally dosed with 100 mg/kg HCBD) were directly infused into the duodenum, approximately 80%
of the biliary metabolites are reabsorbed, with only 20% remaining in the feces and gastrointestinal
tract.
Several studies have reported the identity of excreted metabolites following exposure to 14C-
HCBD. Metabolites identified in the urine of treated rats or mice include £-(1,1,2,3,4-
pentachlorobutadienyl) glutathione, S-(l, 1,2,3,4-pentachlorobutadienyl)-L-cysteine, TV-acetyl-^-
(l,l,2,3,4-pentachlorobutadienyl)-L-cysteine, 1,1,2,3,4-pentachlorobutadienyl sulfenic acid,
1,1,2,3,4-pentachlorobutadiene methyl thioether, 1,1,2,3,4-pentachlorobutadiene
carboxymethylthioether, and 1,1,2,3-tetrachlorobutenoic acid (Dekant et al., 1988a; Nash et al.,
HCBD February 2003 6-7
-------
1984; Reichert and Schultz, 1986; Reichert et al., 1985). As noted previously, the novel metabolite
J/V-acetyl-5'-(l, 1,2,3,4,-pentachlorobutadienyl)-L-cysteine sulfoxide has been detected in the urine
of male, but not female, rats following oral administration of HCBD (Birner et al., 1995).
Comparatively few data are available on the identity of fecal metabolites. Dekant et al.
(1988a) administered a single 30 mg/kg gavage dose of 14C-HCBD in corn oil to male and female
NMRI mice. The feces were collected over a 72-hour period following dose administration.
Approximately 80% of the fecal radioactivity was associated with HCBD. About 10% of the
radiolabel was associated with the HCBD metabolite £-(!,!,2,3,4-pentachlorobutadienyl)
glutathione. The remainder of the fecal radioactivity was present as polar metabolites which could
not be structurally identified.
HCBD February 2003 6-8
-------
7.0 HAZARD IDENTIFICATION
7.1 Human Effects
Limited information is available on the human health effects associated with exposure to
HCBD. A review of the available literature did not identify case reports describing the outcome of
accidental or intentional HCBD exposure, or reports of systemic toxicity following oral or dermal
HCBD exposure. A number of studies have evaluated health effects in workers occupationally
exposed to HCBD via inhalation, and these studies are described below.
7.1.1 Short-Term Studies
No short-term studies describing HCBD health effects in humans were located.
7.1.2 Long-Term and Epidemiological Studies
General Population
No general population studies of HCBD toxicity were located.
Sensitive Populations
No studies concerning HCBD toxicity in sensitive populations were located.
Occupational Exposure Studies
German (1986) conducted two cytogenetic studies of workers employed in an HCBD
production facility. The exposure levels reported by the manufacturer ranged from 1.6 to 16.9
mg/m3. The investigators found an increased frequency of chromosomal aberrations in the peripheral
lymphocytes of exposed workers. However, the frequency of aberrations was not associated with
duration of employment in the HCBD manufacturing facility (WHO, 1994), suggesting that factors
other than HCBD exposure contributed to the observed effects.
Additional occupational studies have evaluated health effects in workers exposed to HCBD.
However, in each case concurrent exposure of workers to other chemicals limits the usefulness of
the data for evaluation of HCBD human health effects. Krasniuk et al. (1969), for example,
evaluated health effects in 153 farm workers intermittently exposed over a period of four years to
soil and grape fumigants containing HCBD. When compared to a control population of 52
unexposed workers, HCBD- exposed workers exhibited increased incidence of arterial hypotension,
myocardial dystrophy, chest pains, upper respiratory tract changes, liver effects, sleep disorders,
hand trembling, nausea, and disordered olfactory functions (U.S. EPA, 1991a). Interpretation of
these data is confounded by concurrent exposure of these workers to polychlorobutane-80.
Burkatskaya et al. (1982) reported adverse health effects in vineyard workers exposed to
fumigants containing HCBD. However, the role of HCBD could not be evaluated because the
workers were concurrently exposed to other agricultural chemicals (WHO, 1994).
HCBD February 2003 1-1
-------
Driscoll et al. (1992) determined the concentrations of individual serum or plasma bile acids
in workers exposed to chlorinated hydrocarbons, including HCBD, carbon tetrachloride, and
perchloroethylene. These investigators reported increases in four serum bile acid parameters in
workers exposed via inhalation to 0.005-0.02 ppm HCBD. The study found no significant relation
between bile acid parameters or liver function tests and exposure. As in the studies above, the
specific contribution of HCBD exposure to the observed effects could not be evaluated.
7.2 Animal Studies
7.2.1 Acute Toxicity
Oral Exposure
Schwetz et al. (1977) reported a single-dose oral LD50 value (the dose that produces lethality
in 50% of the experimental animals) of 65 mg HCBD/kg for male weanling rats and 46 mg/kg for
female weanling rats. A single-dose oral LD50 value of 90 mg/kg was reported for adult rats
(Kennedy and Graepel, 1991). These data suggest that age and gender may be significant variables
in the acute toxicity of HCBD. Single-dose LD50 values reported for other rodents were 80 to 116
mg/kg for mice and 90 mg/kg for guinea pigs (U.S. EPA, 199la).
Three studies have evaluated the non-lethal acute effects of oral HCBD exposure. Nash et
al. (1984) administered a single oral dose of 200 mg/kg HCBD in polyethylene glycol to six male
Wistar-derived rats. Treatment with HCBD increased plasma urea concentration and decreased
plasma alanine aminotransferase activity. Analysis of urine revealed significantly (p > 0.01)
increased levels of glucose, protein, alkaline phosphatase, 7V-acetyl-/-D-glucosaminidase (NAG),
y-glutamyl transpeptidase and alanine aminopeptidase over control levels. These are commonly used
biochemical markers of kidney damage.
Jonker et al. (1993a) investigated the acute oral toxicity of HCBD in 12-week-old male
Wistar rats. The investigators administered single doses of 0, 10, 100, or 200 mg/kg HCBD in corn
oil by gavage to five rats per treatment group. Urine was collected at intervals of 0 to 6 and 6 to 24
hours. All rats were sacrificed at 24 hours. No treatment-related effects were observed at the 10
mg/kg dose. HCBD induced a variety of adverse effects at the two highest dose levels. Kidney
weight, blood plasma creatinine level, urinary pH and occult blood, number of epithelial cells in the
urine, urinary lactate dehydrogenase and NAG activity were significantly increased (p<0.05) at 100
and 200 mg/kg. Additional effects observed in the 200 mg/kg dose group included reduced body
weight, reduced food intake, elevated plasma urea level, and increased urinary volume. Increased
levels of urinary protein, glucose, and potassium, and increased activity of urinary y-
glutamyltransferase and alkaline phosphatase were also observed at 200 mg/kg. Histopathological
examination of the kidneys revealed limited focal necrosis at 100 mg/kg and extensive tubular
necrosis at 200 mg/kg. The study authors identified 10 mg/kg and 100 mg/kg as the acute "No
Nephrotoxic-Effect Level" and "Minimum Nephrotoxic-Effect Level", respectively.
Payan et al. (1993) administered single oral doses of 0, 100, or 200 mg/kg HCBD in
polyethylene glycol to male Sprague-Dawley rats (4 to 5 animals per dose). All rats were sacrificed
HCBD February 2003 7-2
-------
24 hours after exposure, and the right kidneys were subjected to microscopic examination.
Nephrotoxicity was also evaluated by determination of the following biochemical urinary parameters
of renal impairment: urinary concentration of p2-microglobulin (P2-m), urinary yglutamyl
transpeptidase (y-GT) activity, urinary aspartate aminotransferase (AST) activity, and urinary N-
acetyl p-glucosamine (NAG) activity At 100 mg/kg, 4/5 of the rat kidneys exhibited mild or
moderate lesions, compared with 0/4 in the controls. Significant increases in (32-m (2-fold) and AST
(19-fold) were seen at 100 mg/kg, and in all biochemical markers of renal toxicity at 200 mg/kg. A
Lowest-Observed-Adverse-Effect Level (LOAEL) of 100 mg/kg was identified in this study on the
basis of kidney lesions and a increased urinary (32-m and AST excretion.
Lock et al. (1996) administered a single oral dose of 50 mg HCBD/kg to a calf to evaluate
toxic effects on the kidney and bone marrow. The administered dose resulted in the death of the
animal 5 days after treatment. Prior to death, blood urea nitrogen, plasma aspartate aminotransferase
and plasma alkaline phosphatase were elevated, but no changes were observed in circulating white
cells or platelets. The liver and kidneys appeared pale and swollen at necropsy. Histopathological
examination revealed midzonal necrosis in the liver. Extensive areas of necrosis were evident in the
kidney, and were accompanied by hyaline and granular cast formation.
Inhalation Exposure
De Ceaurriz et al. (1988) evaluated the effects of HCBD inhalation exposure on male Swiss
OF1 mice (6 mice/dose). The mice were exposed to HCBD vapor at concentrations between 83 and
246 ppm (886 and 2,625 mg/m3 ) for 15 minutes. Decreased respiratory rates (reflex bradypnea)
were observed at concentrations of 155 ppm (1,652 mg/m3 ) or greater. An EC50 (concentration
producing an effect in 50% of the population) of 211 ppm (2,250 mg/m3) was calculated for this
effect.
De Ceaurriz etal. (1988) investigated the effects of a 4-hour whole-body exposure to HCBD
at measured concentrations of 2.75, 5,10,25 ppm (or 29.3, 53.4,106.7,266.8 mg/m3) on male Swiss
OF 1 mice (10 animals/dose). An HCBD-related increase in the percentage of damaged renal tubules,
as determined by alkaline phosphatase staining, was observed at all exposure levels. The EC50 for
this response was 7.2 ppm (76.8 mg/m3).
Gehring and MacDougall (1971) exposed rats to 161 ppm (1,716 mg/m3) HCBD for 0.88
hour or 34 ppm (362 mg/m3) for 3.3 hours. All rats survived the treatment. Exposure of guinea pigs
or cats under the same conditions resulted in the death of most animals. Inhalation exposure of rats
to 133 to 150 ppm (1,418 to 1,600 mg/m3) for 4 to 7 hours was lethal for some or all animals (NTP,
1991).
Dermal Exposure
Gradiski et al. (1975) evaluated the dermal toxicity and sensitization potential of HCBD in
rabbits. Dermal application of a 10% solution of HCBD (solvent not indicated) to rabbits caused
slight dermal irritation. Guinea pigs exhibited delayed allergic reactions to dermal HCBD
application (U.S. EPA, 1991a).
HCBD February 2003 7-3
-------
Duprat and Gradiski (1978) evaluated the acute toxicity of dermally applied HCBD in female
New Zealand rabbits (10 animals/dose) following an 8-hour exposure period. Undiluted HCBD was
applied at doses of 0.25, 0.5, 0.75 and 1.0 mL/kg under occluded conditions. Four hours after
termination of exposure, the epidermis and subcutaneous tissue revealed edema and
polymorphonuclear leukocyte infiltration at the two highest doses. Three to five days after treatment
at the three highest doses, necrotic changes were noted at the site of application. A few animals from
the two highest dose groups died within 24 hours from respiratory and cardiac failure. Indications
of systemic toxicity included renal epithelial necrosis and fatty liver degeneration. The LD50 for the
eight hour exposure was 0.72 mL/kg (Duprat and Gradiski, 1978). Based on a specific gravity of
1.675 for HCBD (U.S. EPA, 199la), a dermal LD50 of 1,206 mg/kg was calculated.
Acute Ocular Toxicity
Gradiski et al. (1975) reported that instillation of a 10% solution of HCBD (solvent not
reported) into rabbit eyes resulted in slight ocular conjunctival irritation (U.S. EPA, 1991a).
Intraperitoneal Injection
Bai et al. (1992) exposed male Sprague-Dawley rats (4 animals/dose) by intraperitoneal
injection 0, 10.4, 52.2, or 104 mg/kg-day HCBD for three days. Serum bilirubin and alkaline
phosphatase activity was increased (p<0.05) at the two higher doses, indicating disturbance in liver
function. The concentration of total serum bile acids was elevated at the highest dose. No
histopathological examination of the livers was conducted.
Lock and Ishmael (1979) administered single intraperitoneal doses of HCBD in corn oil to
male albino rats (3 to 19 animals/dose). The doses administered were 0, 20, 50, 100, 200, 300, 400,
500, and 1,000 mg/kg. The effects of treatment were assessed 24 hours after administration. Three
of four animals treated with 500 mg/kg died. All rats in the 1,000 mg/kg dose group died. Rats in
the highest dose group exhibited piloerection, sedation, hunching, incoordination, loss of muscle
tone and hypothermia prior to death.
Wolf et al. (1983) conducted studies to examine the effects of HCBD on the cytochrome
P450 content and related monooxygenase activities in the kidneys of male and female rats and male
mice. Adult male and female Wistar rats were administered either 200 or 400 mg/kg HCBD in corn
oil by intraperitoneal injection. Male Swiss-derived mice were administered 50 mg/kg HCBD in
corn oil by intraperitoneal injection. Control animals received corn oil alone (5 ml/kg). A subset of
the animals were treated with 100 mg/kg (3-naphthoflavone (BNF) by intraperitoneal injection daily
for 3 days before HCBD administration. Following HCBD administration, animals were fasted for
24 hours, and renal and hepatic microsomal fractions were obtained. Renal cytochrome P450
concentrations dropped significantly (p < 0.01) in both male and female rats at both doses without
significant losses of cytochrome b5 or NADPH-cytochrome c. In rat renal microsomes, HCBD
increased metabolism of aldrin and 7-ethoxycoumarin significantly (p < 0.05), while metabolism
of p-nitroanisole was only slightly reduced and metabolism of lauric acid was unaffected. No effect
was found on rat hepatic cytochrome P450 levels, and no reductions in rat hepatic monooxygenase
activity were found, although metabolism of 7-ethoxycoumarin was significantly (p < 0.05)
increased. In BNF-treated rats, HCBD decreased renal cytochrome P450 concentrations 33%,
HCBD February 2003 7-4
-------
decreased metabolism of aldrin significantly (p < 0.05), and increased metabolism of 7-
ethoxycoumarin and 7-ethoxyresorufm. In naive mice, HCBD treatment led to an almost complete
loss (15% of control remaining) of cytochrome P450 and associated monooxygenase activity (0-28%
of control metabolism). In BNF-treated mice, a similar loss of cytochrome P450 (17% of control
remaining) resulted in similar decreases in metabolism (0-40% of control metabolism), except for
metabolism of 7-ethoxycoumarin (67% of control metabolism) and 7-ethoxyresorufm (117% of
control metabolism). These findings suggest that the cytochrome P450 subtypes which are decreased
are localized in regions where HCBD damage occurs, while BNF-induced subtypes are not..
Hook and colleagues (Hook et al., 1982, 1983; Kuo and Hook, 1983) conducted a series of
experiments to characterize HCBD toxicity in four different strains of rats. Single doses of 25 to 400
mg/kg HCBD were administered by intraperitoneal injection. Following treatment, relative kidney
weight increased in all dose groups. Organic ion transport was evaluated by analysis of the anion
/7-aminohippurate (PAH) and the cation tetraethylammonium (TEA) in renal cortical slices. Renal
efflux rates of PAH and TEA were unaffected. However, accumulation of the PAH anion was
decreased, while accumulation of the TEA cation was unaffected. This pattern suggests a specific
impact of HCBD on the renal anion uptake system (Hook et al., 1982; Kuo and Hook, 1983).
Kidney-to-body weight ratios increased in all dosage groups. Blood urea nitrogen levels were
elevated in all dose groups, with a more pronounced effect noted in young rats (Kuo and Hook,
1983). Adult male rats were less susceptible to HCBD-induced renal effects than were female adult
rats or young male rats. The investigators attributed this pattern to age- and sex-related differences
in the renal and hepatic enzymes responsible for activation and detoxification of HCBD.
Lock et al. (1984) investigated the effect of age, strain, sex, and monooxygenase inhibitors
on the acute toxicity of HCBD to five different strains of mice (6 or more animals/dose group).
HCBD in corn oil was administered as single intraperitoneal doses ranging from 6.3 to 50 mg/kg.
Toxicity was evaluated 24 hours after treatment. Histopathological examination of the kidneys from
adult Swiss-derived mice of both sexes revealed extensive, dose-dependent proximal tubular
necrosis at doses of 12.5 mg/kg and above. At 6.3 mg/kg, tubular necrosis was only observed
occasionally in a small number of animals. A significant increase in plasma urea occurred in adult
Swiss-derived mice at doses of 25 to 50 mg/kg. No evidence for a gender difference in response was
observed. Young male mice responded to lower doses of HCBD than adults, with an increase in
plasma urea and a decrease in organic ion transport in renal slices evident at 12.5 mg/kg. Prior
administration of the monooxygenase inhibitor piperonyl butoxide or the monooxygenase inducers
phenobarbitone or p-naphthoflavone did not modify the extent of HCBD-induced renal damage.
Intraperitoneal administration of the glutathione or 7V-acetylcy steine conjugates resulted in a pattern
of renal necrosis similar to that observed for HCBD. Evaluation of the comparative susceptibility
of five mouse strains indicated that the BALB/c strain was slightly more susceptible to HCBD
toxicity than the C57BL/10J, C3H, DBA/2J, and Swiss-derived strains.
Ishmael et al. (1984) investigated the time course of histopathology and functional
impairment in adult Swiss-derived mice following a single 50 mg/kg intraperitoneal dose of HCBD
in corn oil. Histopathological examination of the kidneys by light and electron microscopy revealed
mitochondrial swelling within 1 hr after dosing, followed by nuclear pyknosis and increased
cytoplasmic eosinophilia at 4 hr, extensive tubular necrosis after 16 hr, and active tubular
HCBD February 2003 7-5
-------
regeneration within 5 days. Treated animals exhibited significantly increased plasma urea, decreased
renal non-protein sulfhydryl content, and increased renal water content.
7.2.2 Short-Term Studies
Oral exposure
Kociba et al. (1971) conducted a study of HCBD toxicity, the results of which were
published as Kociba et al. (1977) and Schwetz et al. (1977). Female Sprague-Dawley rats (4
animals/dose group) were fed diets containing HCBD at doses of 0, 1, 3, 10, 30, 65, or 100 mg/kg-
day for 30 days. Renal toxicity in the form of increased relative kidney weight as well as renal
tubular degeneration, necrosis and regeneration was observed in rats receiving doses of 30, 65 or
100 mg/kg-day. Minimal hepatocellular swelling was noted at a dose of 100 mg/kg-day. Other
observed effects included decreased food consumption, reduced body weight gain, and increased
hemoglobin concentration at doses of 10, 30,65 or 100 mg/kg-day. No effects were observed in rats
receiving 3 mg/kg-day. This study identified a No-Ob served-Adverse-Effect Level (NOAEL) of 3
mg/kg-day and a LOAEL of 10 mg/kg-day.
Harleman and Seinen (1979) administered diets containing nominal concentrations of 0, 50,
150, or 450 ppm HCBD to weanling Wistar-derived rats (6 animals/sex/dose group) for 14 days.
Based on mean body weight and food consumption data in the study, the mean HCBD doses were
calculated to be 0, 4.6,14.0 and 35.3 mg/kg-day. Body weight and food conversion efficiency were
decreased in a dose-related manner. Food consumption was decreased at 35.3 mg/kg-day. Relative
kidney weights were significantly increased at the two highest dose levels. A dose-related
degeneration of renal tubular epithelial cells was observed in all treated animals, particularly in the
straight limbs (pars recta) of the proximal tubules located in the outer medulla. No indications of
liver toxicity were observed. The low dose of 4.6 mg/kg-day represented the LOAEL.
Stott et al. (1981) conducted an oral exposure study in adult male Sprague-Dawley rats. Five
animals per treatment group were given daily doses of HCBD (0, 0.2 or 20 mg/kg-day) in corn oil
for three weeks. In animals exposed to 20 mg/kg-day, a decrease in body weight gain and an
increase in relative kidney weight were observed. Histopathological examination of the kidneys
revealed damage in the middle and inner cortical regions, with loss of cytoplasm, nuclear pyknosis,
increased basophilia and mitotic activity, and increased cellular debris. No indications of toxicity
were observed in animals exposed to 0.2 mg/kg-day, the NOAEL for this study.
The National Toxicology Program (NTP) conducted a 2-week oral exposure study in B6C3Fj
mice (NTP, 1991; Yang et al., 1989). Groups of mice (5 animal s/sex/dose group) received diets
containing nominal concentrations of 0, 30,100,300,1,000 or 3,000 ppm HCBD for 15 days. Target
concentrations were verified under experimental conditions by gas chromatography. The estimated
daily intake calculated from feed consumption and mean body weights were 0,3,12, and 40 mg/kg-
day for the 0, 30, 100, and 300 ppm dietary concentrations, respectively, for male mice, and 0, 5,
16, and 49 mg/kg-day for female mice. All mice provided with the 1,000 and 3,000 ppm diets died
within seven days. Mice receiving 100 and 3 00 ppm HCBD lost weight. HCBD-related clinical signs
observed at doses of 300 ppm or greater included lethargy, rough hair coats, hunched position, and
incoordination. Marked reductions in thymus and heart weights were noted in mice consuming the
HCBD February 2003 7-6
-------
300 ppm diet. Kidney lesions attributed to HCBD exposure were observed in all treated mice of both
sexes (Yang et al., 1989). Severe necrosis of the cortex and outer medulla was observed in the two
doses that caused deaths among the experimental animals. Necrosis was less severe and
regeneration was prominent in the pars recta of mice receiving lower doses of HCBD.
Histopathologic changes were also observed in liver, lymphoid tissues, and testis of mice in the
1,000 and 3,000 ppm dose groups, but were not clearly related to HCBD toxicity. Minimal-to-mild
depletion of bone marrow in the femur was observed in 2 to 5 mice per dose group in animals
receiving diets containing 300 ppm or higher levels of HCBD. This study identified a LOAEL of
3-5 mg/kg-day in male and female mice, respectively, based on renal tubular necrosis and cellular
regeneration in animals in the lowest dose groups (Yang et al., 1989).
Jonker et al. (1993b) investigated the toxicity of HCBD in 10-week-old male and female
Wistar rats (5 rats/sex/dose). HCBD was provided in the diet at levels of 0, 25, 100, or 400 ppm for
a duration of four weeks. Based on mean body weight and food intake data in the study, these
concentrations correspond to average daily doses of 0, 2.25, 8, and 28 mg/kg-day. Treatment-
induced signs of toxicity were observed at doses of 8 and 28 mg/kg-day in both sexes. The observed
signs included decreased liver weight, tubular cytomegaly, decreased plasma creatinine, decreased
body weight (10% in males and 15% in females), and decreased adrenal weight. Increased plasma
aspartate aminotransferase activity and bilirubin were observed at the 28 mg/kg-day dose. The
NOAEL and LOAEL identified from this study were 2.25 and 8 mg/kg-day, respectively.
Lock et al. (1996) conducted two short-term experiments on the effects of HCBD in calves.
Each experiment evaluated toxicity in a single animal. In the first experiment, a dose of 5 mg/kg-day
was administered orally for 7 days. An increase in blood urea nitrogen was noted after the fifth dose,
and levels remained high until the animal was euthanized nine days after initiation of treatment.
Plasma levels of aspartate transaminase and alanine aminotransferase were elevated, but no changes
were observed in hematological parameters. At necropsy, perirenal edema was observed in the
kidneys and the liver was pale and swollen. Histopathological examination revealed midzonal
necrosis of the liver and extensive swelling of the tubular epithelium and degenerative changes in
the kidney. Casts were evident in the tubules of the medulla.
In the second short-term experiment conducted by Lock et al. (1996), a calf was dosed with
2.5 mg HCBD/kg-day for 10 days and the blood was monitored for 20 days for urea and platelet
count. The dose was subsequently increased to 5 mg/kg-day, with 8 doses administered over 12
days. A marginal increase in blood urea nitrogen was observed on day 14. Aspartate transaminase
and alanine aminotransferase were increased on day 7, and gradually decreased to normal levels on
day 15. The calf was euthanized and necropsied 18 days after the start of the dosing regimen.
Histopathological examination revealed slight disruption of the midzonal architecture of the liver,
while mild renal tubular degeneration was evident in the kidney. The results of the experiments
conducted by Lock et al. (1996) indicate that HCBD is both a nephro- and hepato-toxicant in calves.
Nakagawa et al. (1998) exposed male Wistar rats (3 animals/dose group) to 0, 0.008, 0.04,
or 0.2% HCBD in the diet for 3 weeks. Assuming a food consumption factor of 0.09 kg/kg/day (U. S.
EPA, 1988), these dietary levels correspond to approximate daily doses of 0, 7.2, 36, and 180
mg/kg-day. Rats ingesting the 0.04% and 0.2% diets had lower mean body weight (decreases of 15%
and 46%, respectively) at the termination of the experiment. Kidney weight was unaffected.
HCBD February 2003 7-7
-------
Histopathological examination of rats in the 180 mg/kg-day group revealed indications of extensive
regeneration in the straight portion (pars recta) of the proximal tubule. Similar lesions were not
evident at lower doses. These data suggest a NOAEL of 7.2 mg/kg-day based on absence of effect
on weight gain or renal histopathology.
Inhalation exposure
NIOSH (1981) reported 100% mortality in mice exposed to HCBD vapors for 5 days, 7
hours/day, at a concentration of 50 ppm (533 mg/m3), but no deaths in animals exposed to 10 ppm
(106.6 mg/m3).
Gage (1970) conducted an inhalation study in Alderley Park SPF rats. Groups of adult rats
(4/sex/dose) were exposed to nominal HCBD concentrations of 53, 107, or 267 mg/m3 for 15 days,
6 hours/day (duration-adjusted concentrations of 13, 27, or 67 mg/m3); 1,067 mg/m3 for 12 days, 6
hours/day (267 mg/m3 duration-adjusted); or 2,668 mg/m3 for 2 days, 4 hours/day (445 mg/m3
duration-adjusted). Petroleum ether was used as a solvent for concentrations below 1,067 mg/m3.
No indications of toxicity were observed at the lowest level of exposure, suggesting a NOAEL of
53 mg/m3 (13 mg/m3 duration-adjusted). Two of the four female rats exposed to 1,067 mg/m3 HCBD
died. Pale enlarged kidneys, adrenal regeneration, and renal cortical tubular degeneration with
epithelial regeneration were noted at autopsy. Surviving females at this concentration were slightly
anemic. The weight gain of female rats was reduced at 107 and 267 mg/m3. At 1,067 mg/m3, rats of
both sexes lost weight. Irritation of the eyes and nose was observed at the two highest levels of
exposure. Respiratory distress occurred at concentrations of 267 mg/m3 or greater. At the
termination of the experiment, enlarged pale kidneys were evident in the 267 and 1,067 mg/m3
treatment groups. Enlarged adrenals were observed in animals exposed to 1,067 mg/m3.
Histopathological analysis revealed degeneration in the adrenal cortex and proximal tubules of the
kidneys at concentrations of 267 mg/m3 or greater (WHO, 1994).
7.2.3 Subchronic Studies
Schwetz et al. (1977) fed male (10 -17 per dose group) and female (20 - 34 per dose group)
Sprague-Dawley rats a diet containing 0.2, 2.0, or 20 mg/kg-day HCBD for evaluation of
reproductive effects. HCBD was provided in the diet before and during mating, and throughout
gestation and lactation, for a total study duration of 148 days. Adult rats from the 20 mg/kg-day dose
level had decreased body weight gain along with decreased food consumption. At necropsy, relative
kidney weights were increased in high-dose males and females. Relative liver weight was increased
in high-dose males, and relative brain weight was increased in high-dose females. The kidneys of
males exposed to 2 (1 of 10 examined) or 20 (3 of 15 examined) mg/kg-day were roughened and
had a mottled cortex. Histopathological examination revealed dose-related increases in tubular
dilation and regeneration in animals exposed to 2 or 20 mg/kg-day (Table 7-1). These results
indicate a NOAEL of 0.2 mg/kg-day for male and female rats, based on the absence of observed
renal histopathology or other toxic effects at this dose.
HCBD February 2003 7-8
-------
Table 7-1. Histopathological Findings in Adult Rats Fed Diets containing
Hexachlorobutadiene.
Administered Dose
(mg/kg-day)
Moderate Proteinaceous Casts
in Dilated Renal Tubules
Moderate Focal Renal Tubular
Collapse and Atrophy
Dilation and Hypertrophy of the
Tubules in the Kidney Cortex
Male Rats
0
1/5
1/5
1/5
0.2
0/5
0/5
0/5
2
3/5
3/5
0/5
20
4/5
4/5
4/5
Female Rats
0
0/5
0/5
0/5
0.2
0/5
0/5
0/5
2
1/5
4/5
1/5
20
0/5
0/5
4/5
source: Schwetz etal. (1977)
Harleman and Seinen (1979) exposed groups of weanling Wistar-derived rats (10/sex/dose
group) to daily oral gavage doses of 0, 0.4, 1.0, 2.5, 6.3, or 15.6 mg/kg-day HCBD in a peanut oil
vehicle. The study duration was 13 weeks. Reductions in body weight gain, food consumption, and
food utilization efficiency were noted in the two highest dose groups. Dose-related increases in
relative kidney weights were noted in all treatment groups of male mice, and in females administered
6.3 or 15.6 mg/kg-day. Proximal tubular degeneration was noted in males treated with doses of 6.3
mg/kg-day and above, and in females treated with doses of 2.5 mg/kg-day and above. This effect
was characterized by hyperchromatic nuclei, hypercellularity, vacuolation and focal necrosis of renal
epithelial cells, and a diminished brush border. Polyuria and decreased urine osmolarity were noted
in females receiving doses equal to or greater than 2.5 mg/kg-day, and in males receiving 15.6
mg/kg-day. Relative liver weights were increased in females at 15.6 mg/kg-day and in males at 6.3
and 15.6 mg/kg-day. Histological examination of the liver revealed increased cytoplasmicbasophilia
only in males treated with 6.3 mg/kg-day and above. Relative spleen weights were increased in
males at 15.6 mg/kg-day, and in females at the two highest doses. The study authors identified
NOAEL values of 1.0 mg/kg-day for females and 2.5 mg/kg-day for males.
NTP (1991) conducted a 13-week oral exposure study in B6C3Fj mice. Groups of mice (10
animals/sex/dose group) received diets containing 0, 1, 3, 10, 30, or 100 ppm HCBD. Target
concentrations were verified under experimental conditions by gas chromatographic analysis. Based
on average food consumption and body weight data, these concentrations corresponded to dose
levels of 0, 0.1, 0.4, 1.5, 4.9 or 16.8 mg/kg-day for males and 0, 0.2, 0.5, 1.8, 4.5 or 19.2 mg/kg-day
for females. No HCBD-related clinical signs or deaths were observed. Food consumption of treated
and control animals was similar. Reduced body weight gain was reported in males exposed to diets
containing 30 or 100 ppm HCBD (decreases of 49% and 56%, respectively) and in females exposed
to the 100 ppm diet (47%). Relative kidney weight was significantly decreased (p<0.01) in the three
highest dose groups of males, and in females in the highest dose group. High-dose males also
exhibited decreased relative heart weight, although no histologic lesions were reported in this organ.
HCBD February 2003
7-9
-------
Histopathological changes were noted in the kidneys of treated animals. Necropsy revealed
treatment-related increases in renal tubular cell regeneration. This lesion was characterized as a
diffuse increase in epithelial nuclei and increased basophilic staining. Female mice appeared to be
more susceptible than male mice to the formation of this lesion following exposure to HCBD, with
occurrence noted at dose levels of 0.2 mg/kg-day and above. Incidence of regeneration in female
mice was 0% at control, 10% at 0.2 mg/kg-day, 90% at 0.6 mg/kg-day, and 100% at higher doses.
Lesions were observed in male mice at dose levels of 4.9 mg/kg-day and above, with incidences of
100% in these groups (NTP, 1991; Yang et al., 1989). In contrast to results from the 2-week study
conducted by the same investigators, no evidence of necrosis was observed. Based on the
histopathologic evaluation of the kidney, the authors identified aNOAEL of 1.5 mg/kg-day for male
mice. Because tubular regeneration occurred in 1 of 10 females in the lowest dose group (0.2 mg/kg-
day), the study authors concluded that aNOAEL for female mice could not be identified from these
data (NTP, 1991). However, others have concluded that the effect observed at 0.2 mg/kg-day is not
statistically significant, and therefore considered this dose to be the NOAEL for female mice (WHO,
1994; U.S. EPA, 1998a). Further statistical analysis shows that, although the 0.2 mg/kg-day dose
fails a Fisher's exact test against control (p = 0.5), the data passes a Mantel-Haenszel trend test (p
< 0.001), suggesting that a benchmark dose analysis would be appropriate.
Nakagawa et al. (1998) administered 0.1% HCBD in the diet to male Wistar rats (21 rats/
group) for 30 weeks in conjunction with a cancer promotion study (discussed in Section 7.2.7).
Assuming a food consumption factor of 0.09 kg/kg/day (U.S. EPA, 1988), this dietary level
corresponds to an average daily dose of 90 mg/kg-day. HCBD treatment resulted in decreased final
body weight, and increased relative kidney weight. No significant differences were noted in serum
and urine biochemical parameters. Simple hyperplasia of renal tubular structures was observed, but
the incidence did not differ significantly from the control. Histopathological examination did not
reveal adenomatous hyperplastic foci or renal tumors.
7.2.4 Neurotoxicity
Data from distribution studies indicate that HCBD accumulates in brain tissue (Reichert et
al., 1985). This observation raises the possibility that HCBD exposure may affect neurological
function. Studies designed to specifically evaluate neurotoxicological endpoints following HCBD
exposure were not identified in the available literature. However, neurological effects have been
observed in a number of oral and dermal exposure studies. Kociba et al. (1977) reported increased
relative brain weights in female rats fed 20 mg/kg-day HCBD for 2 years. This increase occurred
in a dose group with depressed body weights and was not accompanied by histopathological changes
in the brain. Similarly, Schwetz et al. (1977) noted depressed body weight and increased relative
brain weight in female rats fed 20 mg/kg-day for 148 days in a reproductive study. Concurrent
changes in behavior or brain histopathology were not observed in the affected animals. An increase
in relative brain weights with decreased body weights was also observed in male and female B6C3Fj
mice fed 16.8 to 19.2 mg/kg-day HCBD in their diet in a 90-day subchronic study (NTP, 1991).
Treatment-associated neurotoxic effects were observed by Harleman and Seinen (1979), who
provided female Wistar rats (6 animals/dose) with diets containing 0, 150, or 1,500 ppm HCBD
(corresponding to average daily doses of 0,15 or 150 mg/kg-day). The duration of exposure ranged
from 10 to 18 weeks. Indications of neurotoxicity observed at the 150 mg/kg-day dose included
HCBD February 2003 7-10
-------
ataxia, incoordination, weakness of the hind legs, and unsteady gait. Histopathological examination
revealed demyelination and fragmentation of femoral nerve fibers in high-dose animals. No
treatment-related histopathological changes were observed in the brain.
In Russian studies cited through secondary sources (WHO, 1994), Badaeva (1983) and
Badaeva et al. (1985) observed that daily oral administration of 8.1 mg/kg-day HCBD to pregnant
rats throughout gestation resulted in histopathological changes in nerve cells and myelin fibers of
the spinal cord in the dams and their offspring. Increased levels of free radicals were detected in the
brain and spinal cord of the offspring of treated dams (U.S. EPA, 199 la).
Neurotoxicity has also been observed following HCBD exposure by the dermal route. Duprat
and Gradiski (1978) observed central nervous system depression manifested as stupor in rabbits
following application of 0.25 to 1.0 ml/kg (418 to 1,675 mg/kg) in an acute dermal toxicity test.
Stupor was observed throughout the 8-hour exposure period, and during a 2-hour period
immediately following exposure.
7.2.5 Developmental/Reproductive Toxicity
Oral Exposure
Schwetz et al. (1977) provided male and female Sprague-Dawley rats (30 to 51 animals/dose
group) with a diet containing HCBD at levels corresponding to doses of 0, 0.2, 2.0, or 20 mg/kg-day
HCBD. The HCBD-containing diet was administered for 90 days prior to mating, 15 days during
mating, 22 days during gestation, and 21 days during lactation. Adults in the two higher dose groups
exhibited multiple signs of toxicity, including decreased food consumption, decreased body weight
gain, and renal tubular degeneration. No HCBD-related effects on pregnancy rate, time to delivery,
neonatal survival, neonate sex ratio, weanling histopathology, or incidence of neonatal external,
visceral, or skeletal anomalies were observed. Slightly decreased pup weight (p<0.05) was observed
in the 20 mg/kg-day treatment group at postnatal day 21. The identified NOAEL and LOAEL for
developmental effect were 2 and 20 mg/kg-day, respectively.
Harleman and Seinen (1979) provided female rats (6 animals/dose) with a diet containing
HCBD at levels of 0, 150, or 1,500 ppm for 3 weeks prior to mating, 3 weeks during mating, and
throughout gestation and lactation. Assuming a food consumption factor of 0.1 kg/kg/day (U.S.
EPA, 1988), these concentrations correspond to average daily doses of 0, 15, or 150 mg/kg-day.
High-dose females were sacrificed at week 10 of the study, and low-dose females were sacrificed
at week 18. Maternal toxicity occurred in both dose groups, and observed effects included reduced
body weight gain, increased relative kidney weight, and histopathological changes in kidneys.
Neurological effects, including ataxia, incoordination, weakness of the hind legs, and unsteady gait,
were observed in the dams at the 150 mg/kg-day dose. Furthermore, at the 150 mg/kg-day dose no
conceptions occurred, the ovaries showed little follicular activity, and no uterine implantation sites
were observed. At 15 mg/kg-day, fertility and litter size were reduced, but the effect was not
statistically significant. Pup weights in this treatment group were significantly reduced on postnatal
days 0, 10, and 20. No gross abnormalities were noted. The LOAEL identified was 15 mg/kg-day.
HCBD February 2003 7-11
-------
In Russian studies cited in a secondary source (WHO, 1994), Badaeva and colleagues
(Badaeva, 1983; Badaeva et al., 1985) conducted two studies in which pregnant female rats were
orally administered 8.1 mg/kg-day HCBD throughout gestation. Offspring of HCBD-treated dams
had lower body weight and shorter crown-rump length when compared to controls (U.S. EPA,
199 la). Histological changes in the nerve cells and myelin fibers of the spinal cord were noted in
both dams and offspring. Neurological changes reported in the offspring included ultrastructural
changes in neurocytes and increased levels of free radicals in the brain and spinal cord (U.S. EPA,
199 la).
In addition to the reproductive and developmental toxicity studies discussed above, two
longer-term toxicity studies have evaluated reproductive endpoints following oral exposure to
HCBD. No treatment-related lesions in reproductive organs were observed in rats that received
lifetime exposures of up to 20 mg/kg-day HCBD (Kociba et al., 1977). No significant changes were
noted in sperm count, the incidence of abnormal sperm, estrual cyclicity, or the average estrous
cycle length in mice administered 100 ppm HCBD in the diet for 13 weeks. Sperm motility in treated
mice was significantly lower, though not in a dose-related manner, than that observed for controls
(NTP, 1991).
Inhalation Exposure
Saillenfait et al. (1989) exposed pregnant Sprague-Dawley rats (24 animals/dose) to HCBD
vapor at nominal concentrations of 0, 21, 53, 107, or 160 mg/m3 (0, 2, 5, 10, or 15 ppm) for 6
hours/day from gestation days (GD) 6-20, resulting in duration-adjusted concentrations1 of 0, 5,13,
27, or 40 mg/m3. Concentrations were monitored by gas chromatography. Animals were sacrificed
on GD 21. Decreased body weight gain was occurred in dams exposed to 53 or 160 mg/m3. Body
weight was decreased (p<0.01) in male (9.5%) and female (12.9%) fetuses in the 160 mg/m3
treatment group. Fetal body weight was unaffected at lower doses. The mean number of implantation
sites, total fetal losses, resorptions, number of live fetuses, pregnancy rate, and fetal sex ratio were
comparable in the treated and control groups. No exposure-related external, visceral, or skeletal
anomalies were noted in any dose group. The NOAEL for maternal toxicity is at 5 mg/m3.
In dominant lethal tests in CD (Sprague-Dawley-derived) rats, exposure to HCBD vapors
at 10 or 50 ppm (107 or 533 mg/m3) for 5 consecutive days, 7 hours/day, did not affect fertility
indices, number of corpora lutea or implantations, or the frequency of early death (NIOSH, 1981).
For B6C3FJ mice that were exposed to HCBD vapors at 107 or 533 mg/m3, all animals in the
high-dose group died during the 5-week post-treatment period (NIOSH, 1981). The frequency of
abnormal sperm morphology in the low-dose group did not differ significantly from controls.
Although the Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA,1991b) recommended
against dosimetric adjustment of developmental toxicity data, the draft Review of the Reference Dose and
Reference Concentration Processes (U.S. EPA, 2002) recommends that duration adjustment procedures to
continuous exposures be used for inhalation developmental toxicity studies as for other health effects from
inhalation exposure.
HCBD February 2003 7-12
-------
Intraperitoneal Injection
Mated female Sprague-Dawley rats (10-15 animals/group) received 10 mg/kg-day HCBD
in corn oil via intraperitoneal injection, during gestation days 1 to 15 (Hardinetal., 1981). Maternal
toxicity consisted of changes in two organ weights (no further details provided). Decreased pre- and
post-implantation survival was also noted. Developmental effects included decreased fetal weight
or length, a l-to-2-day delay in heart development, and dilated ureters. Gross external and internal
examinations revealed no malformations (WHO, 1994).
Harris et al. (1979) exposed pregnant female rats to 10 mg/kg-day HCBD from gestation
days 1 to 15 via intraperitoneal injection. A 3-fold increase in soft tissue anomalies was reported in
offspring of treated dams. No particular type of anomaly predominated (U.S. EPA, 199 la).
7.2.6 Chronic Toxicity
Data from a single chronic oral exposure study are available in the published literature.
Kociba et al. (1977) provided male and female Sprague-Dawley rats (39 to 40/sex/dose level; 90/sex
for controls) with diets that contained 0, 0.2, 2, or 20 mg/kg-day HCBD (99% pure) for 22 months
(males) or 24 months (females). Parameters monitored included appearance and demeanor, body and
organ weights, food consumption, hematologic and urine analysis, urinary porphyrins, serum
clinical chemistry, and histopathology of major organs. The investigators reported significantly
increased mortality in high-dose males (p<0.05), but incidences were not given. Decreased body
weight gain was noted in high-dose males and females, with significant differences (p<0.05) from
controls evident after day 27 (females) or day 69 (males). There were no apparent treatment-related
effects on food consumption. High-dose animals had increased relative brain weights (females) and
relative testes weights (males).
An important observation in the Kociba et al. (1977) study is the clear dose-response
relationship for HCBD-induced renal toxicity. No discernible effects were noted at the 0.2 mg/kg-
day dose. Effects noted at the intermediate dose of 2 mg/kg-day included increased urinary
coproporphyrin excretion (females only, days 427-428), and increased renal tubular epithelial
hyperplasia. Lifetime ingestion of the 20 mg/kg-day dose resulted in increased urinary excretion of
coproporphyrin and increased terminal weight of the kidneys in rats of both sexes (females - days
728-729, males - days 377-378). The physiological relationship of a change in coproporphyrin
excretion to renal toxicity is not currently clear. Microscopic examination revealed histopathological
changes in the kidney, including hyperplasia and neoplasia of the renal epithelium. No incidence
data was provided for the non-neoplastic effects. HCBD-related neoplastic changes are further
discussed in Section 7.2.7. The lowest dose of 0.2 mg/kg-day was identified as the NOAEL in this
study. The LOAEL was 2 mg/kg-day.
7.2.7 Carcinogenicity
Oral Exposure
Kociba et al. (1977) observed the tumorigenic potential of HCBD in male and female rats
fed 0, 0.2, 2, or 20 mg/kg-day in a 2-year oral exposure study. No adverse effects attributable to
HCBD February 2003 7-13
-------
HCBD were observed in the low-dose group. Ingestion of the intermediate 2 mg/kg-day dose
resulted in signs of renal tubular epithelial hyperplasia, but no evidence of neoplasia was observed.
Ingestion of 20 mg/kg-day for 2 years resulted in development of renal tubular adenomas and
adenocarcinomas. Neoplasms were observed in approximately 23% (9/39) of the males and 15%
(6/40) of the females at 20 mg/kg-day, compared with 1.1% (1/90) and 0% (0/90) in control male
and female rats and 0% at 0.2 or 2 mg/kg-day. Of the observed neoplasms, 7 in the high dose males
and 3 in the high dose females were malignant, the rest being benign. Combined incidence of renal
tubular benign and malignant tumors was significantly increased when compared to controls
(p<0.05) for both males and females. Metastasis to the lungs was observed in two of the treated
animals. An important observation in this study was that HCBD-induced neoplasms occurred only
at a dosage level that caused substantial renal tissue injury. Additional details of this study are
provided in Section 7.2.6.
In a Russian study cited through secondary sources (USEPA, 1991a), Chudin et al. (1985)
conducted an HCBD oral exposure study in male Wistar rats for approximately 1 year. The doses
of HCBD, administered by gavage in sunflower oil, were 0.6, 5.8 or 37 mg/kg-day (n=43, 43, and
41, respectively). Control rats were either untreated (n=90) or received only the sunflower oil
vehicle (n=46). Some benign tumors of the kidney and liver were reported.
Nakagawa et al. (1998) investigated the effect of HCBD on renal carcinogenesis in male
Wistar rats pre-treated with TV-ethyl-7V-hydroxyethylnitrosamine (EHEN). EHEN is a known
nephrocarcinogen in rats, where it selectively induces renal tubular cell tumors. The purpose of this
study was to evaluate the ability of HCBD to act as a promoting stimulus following subthreshold
exposure to EHEN. HCBD was administered for 30 weeks at a concentration of 0.1% by weight
(dose calculated to bel,000 mg/kg) in the diet of rats (12/treatment group) that had previously
received 0.1% EHEN in the drinking water. The combined treatment with HCBD and EHEN
resulted in a significantly higher renal tumor incidence than did administration of EHEN alone. Rats
treated with HCBD alone did not develop renal tumors under the conditions used in this
investigation. Significantly increased levels of bromodeoxyuridine (BrdU) labeling indicated
increased cell proliferation in the outer stripe and cortex of kidneys from HCBD-treated rats. In a
parallel experiment, immunostaining for proliferating cell nuclear antigen (PCNA) was used to
estimate nuclear DNA synthesis in defined renal tubular segments of HCBD-treated rats. A
significant increase in the number of PCNA-positive cells was noted only in the outer stripe. These
results are consistent with the outer stripe as a site for renal lesions induced by HCBD. Nakagawa
et al. (1998) concluded that the ability of HCBD to induce EHEN-initiated carcinogenesis appears
to be associated with nephropathy and subsequent cell proliferation.
Dermal Exposure
Van Duuren et al. (1979) evaluated the carcinogenicity of dermally applied HCBD in female
Swiss mice (30 animals/group). The investigators applied 6.0 mg of HCBD in acetone to shaved
dorsal skin three times per week for a duration of 440 to 594 days. The treatment did not increase
the incidence of papillomas or carcinoma at the site of application, or the incidence of tumors at
distant sites such as the lung, stomach or kidney.
HCBD February 2003 7-14
-------
Van Duuren et al. (1979) also evaluated HCBD in an initiation-promotion experiment.
Female Swiss mice (30 animals/group) received a single application of 15.0 mg HCBD in acetone
on shaved dorsal skin. Fourteen days after HCBD application, dermal applications of 5 jig of the
tumor promoter 12-o-tetradecanoylphorbol-13-acetate (TPA) were administered to the test site three
times per week for a total duration of 428 to 576 days. The incidence of skin papillomas in HCBD-
treated animals was comparable to that in controls.
Intraperitoneal Injection
Theiss et al. (1977) investigated the carcinogenic potential of HCBD by assessment of the
pulmonary tumor response in male strain A/St mice. Twenty animals per dose group were given
intraperitoneal injections of 4 or 8 mg/kg HCBD in tricaprylin, three times per week for a total of
13 and 12 injections, respectively. The total injected dose was 52 or 96 mg HCBD per animal. All
surviving animals were killed 24 weeks after the first injection, and were examined for pulmonary
surface adenomas. The tumor incidences were similar in treated and control groups. However, the
use of this study for the evaluation of the carcinogenicity of HCBD is limited by the use of a mouse
strain that is highly predisposed to spontaneous lung cancer, the small number of animals per dose
group, the parenteral route of administration, the limited scope of histopathological evaluation
(WHO, 1994), and short study duration (6 months).
7.3 Other Key Data
7.3.1 Mutagenicity/Genotoxicity
The mutagenicity of HCBD has been evaluated in an array of in vivo and in vitro assays. The
results of these tests are summarized below by category. No information was located regarding the
genotoxic effects of HCBD in humans.
Bacterial Test Systems
Test results from bacterial assays of mutagenicity are summarized in Table 7-2. Most tests
in standard S. typhimurium reverse mutation assays have been negative, with or without S9
activation (Rapson et al., 1980; Reichert et al., 1983; Stott et al., 1981; DeMeester et al., 1981;
Haworth et al., 1983; Chudin et al., 1985) except for a study by Simmon (1977) in which a positive
response in Salmonella typhimurium was reported in the presence of metabolic activation induced
by rat liver S9 fraction (U.S. EPA, 1991a). Conflicting results in standard assays may be due to
contaminants in technical and even analytical grade HCBD (Reichert et al., 1984; Vamvakas et al.,
1988). Vamvakas et al. (1988) observed 98% pure HCBD was a direct-acting mutagen in S.
typhimurium TA 100; but after HCBD was purified to 99.5%, a negative mutagenic response was
obtained.
Positive results have been reported for HCBD when pre-incubated with both rat liver
microsomes and glutathione, but not when either was omitted (Vamvakas et al., 1988; Roldan-
Arjona et al., 1991). The mutagenic response was increased by additional inclusion of rat kidney
microsomes, mitochondria, or cytosol as yglutamyl transpeptidase and dipeptidase sources, and
reduced by addition of a (3-lyase inhibitor (Vamvakas et al., 1988). Reichert et al. (1984) also
HCBD February 2003 7-15
-------
reported a positive response in S. typhimurium when a "fortified" S9 mix, containing 3-fold more
S9 protein than standard Ames test protocols, was utilized.
Test results for mutagenicity assays of HCBD metabolites are summarized in Table 7-3.
Positive results in bacterial reverse mutation assays have been obtained for the mono-glutathione
and mono-cysteine conjugates of HCBD (Green and Odum, 1985; Dekant et al., 1986; Vamvakas
et al., 1988) and the mercapturic acid, TV-acetyl-S-pentachlorobutadienyl-L-cysteine (Reichert and
Schutz, 1986; Wild et al., 1986). Other potential HCBD metabolites that gave positive results in
reverse mutation assays were pentachloro-3-butenoic acid and pentachloro-3-butenoic acid chloride
(Reichert et al., 1984), in the absence of S9 activation. When rat kidney fractions were used for
metabolic activation, the addition of a specific inhibitor of P-lyase (aminooxyacetic acid) to the
system reduced the mutagenic response (Vamvakas et al., 1988), indicating that HCBD metabolites
mediated mutagenesis in these assays. These results are consistent with the proposed mechanism for
bioactivation of HCBD in animals (Figure 6-1, Section 6.3).
In Vitro Mammalian Cell Test Systems
Data for mutagenicity assays in mammalian test systems are summarized in the upper portion
of Table 7-4. Treatment with HCBD did not increase the frequency of chromosome aberrations in
Chinese hamster ovary (CHO) cells (Galloway et al., 1987) or cultured human lymphocytes
(German, 1988). However, Galloway et al. (1987) observed a significant increase in sister chromatid
exchange in CHO cells treated with HCBD. Schiffman et al. (1984) reported unscheduled DNA
synthesis (UDS) activity and morphological transformation in Syrian hamster embryo fibroblasts
with and without metabolic activation. Stott et al. (1981) reported negative results in a rat primary
hepatocyte UDS assay.
Vamvakas et al. (1989) evaluated the genotoxicity of the HCBD metabolite S-( 1,2,3,4,4-
pentachlorobutadienyl)glutathione (PCBG) in cultured porcine kidney LLC-PK cells. Incubation
of confluent monolayers with PCBG resulted in a dose-dependent induction of DNA repair. Addition
of either acivicin, an inhibitor of y-glutamyl transpeptidase, or aminooxyacetic acid, an inhibitor of
cysteine conjugate P-lyase, prevented PCBG-induced genotoxicity. These results are consistent with
the hypothesis that renal metabolism plays a key role in PCBG-induced genotoxicity.
HCBD February 2003 7-16
-------
Table 7-2. Mutagenicity of HCBD in Salmonella typhimurium Test Systems.
Strain
TA100
TA 1535
TA100
TA100
TA98
TA100
TA1530
TA1535
TA1538
TA100
TA98
TA100
TA98
TA100
TA1535
TA1537
TA100
TA1535
TA1538
TA100
TA100
TA100
Conditions
-
purity not reported
purity >99%
purity 98%, plate incorporation
purity > 99%, plate incorporation
purity not reported, suspension test
purity not reported, preincubation test
+ fortified S9*a,
purity > 99.5%, preincubation test
+/- rat liver S9, purity not reported, plate
incorporation
+/- rat liver S9, preincubation test
+ rat liver S9*, preincubation test
+ rat liver S9*, purity 99.5%, preincubation
test
purity 99.5%
Results"
With S9 Activation
+
nd
-
-
-
-
+
-
-
+
+c
nd
Without S9 Activation
+
-
-
-
-
-
nd
-
-
-
nd
+
Reference
Simmon (1977)
Rapsonetal. (1980)
Stottetal. (1981)
De Meester et al. (1981)
Stottetal. (1981)
Reichert et al. (1983)
Haworthetal. (1983)
Chudinetal. (1985)
Reichert etal. (1984)
Wild etal. (1986)
Vamvakas etal. (1988)
HCBD February 2003
7-17
-------
Table 7-2 (continued)
Strain
TA100
TA100
Conditions
purity 99.5%, preincubation test
+ rat liver microsomes without additional
GSH
+ rat liver microsomes and additional GSH,
purity 99.5%,
plate incorporation
purity 98%, preincubation test
Results"
With S9 Activation
nd
-
+d
-
Without S9 Activation
-
nd
nd
+
Reference
Vamvakasetal. (1988)
Roldan-Arjona et al. (1991)
Source: adapted from WHO (1994)
a S9* = a fortified S9 mix containing 3 times the normal protein concentration; GSH = reduced glutathione
b + = > twice the background rate or, in the case of bacterial studies, a reproducible dose-related increase in the number of revertants per plate; - = negative; nd
= not determined
°0.23 revertants per nmol
d Addition of rat kidney microsomes further increased the number of revertants; positive results were inhibited by the p-lyase inhibitor
aminooxyacetic acid
HCBD February 2003
7-18
-------
Table 7-3. Mutagenicity of HCBD Metabolites.
Metabolite3
PCBG
TCBG
TCBC
PCBC
N-AcPCBC
PCCMT
PCMT
PCBA
PCBAC
Conditions
no activation
+ rat kidney S9
+/- rat kidney fractions
+/- rat kidney fractions
+/- rat kidney fractions
+/- rat kidney S9
no activation
- rat liver S9
+ rat liver S9
+ rat liver S9
+ rat liver S9*b
+ rat liver S9
+ rat liver S9
+ rat liver S9
Result0
With S9
Activation
nd
+
+d
-
-
+e
nd
nd
+
+
-
-
+
+
Without S9
Activation
-
nd
+d
-
-
+e
+f
-
nd
nd
nd
nd
+
+
Reference
Green & Odum (1985)
Green & Odum (1985)
Vamvakasetal. (1988)
Vamvakasetal. (1988)
Vamvakasetal. (1988)
Green & Odum (1985)
Dekantetal. (1986)
Wild etal. (1986)
Reichert and Schutz
(1986)
Wild etal. (1986)
Wild etal. (1986)
Reichert etal. (1984)
Reichert etal. (1984)
Source: modified from WHO (1994)
Abbreviations: PCBG S-(l,l,2,3,4-pentachlorobutadienyl) glutathione
TCBG 1,4 bis(l,2,3,4-tetrachlorobutadienyl) glutathione
TCBC 1,4 bis (l,2,3,4-tetrachlorobutadienyl)-L-cysteine
PCBC S-( 1,1,2,3,4-pentachlorobutadienyl)-L-cy steine
jV-AcPCBC N-acetyl-S-( 1,1,2,3,4-pentachlorobutadienyl)-L-cysteine
PCCMT 1,1,2,3,4-pentachlorobutadiene carboxymethylthioether
PCMT 1,1,2,3,4-pentachlorobutadiene methylthioether
PCBA 2,2,3,4,4-pentachloro-3-butenoic acid
PCBAC 2,2,3,4,4-pentachloro-3-butenoic acid chloride
b S9* = a fortified S9 mix containing 3 times the normal protein concentration; GSH = reduced glutathione
0 + = twice background rate; - = negative; nd = not detectable
d Mutagenic potency enhanced by rat kidney microsomes or mitochondria and less so by cytosol; positive results
e Mutagenic potency enhanced by rat kidney microsomes or mitochondria and less so by cytosol; positive results
f Mutagenic potency enhanced by the (3-tyase inhibitor aminooxyacetic acid
HCBD February 2003
7-19
-------
Table 7-4. Genotoxicity of HCBD in Eukaryotic Assay Systems.
Species/Strain/
Cell Type
Compound
End Point
Comments
Results3
Reference
In Vitro Assays
Chinese hamster
ovary cells
Human lymphocytes
Chinese hamster
ovary cells
Syrian hamster
embryo fibroblast
Porcine kidney cells
Syrian hamster
embryo fibroblasts
Rat primary
hepatocyte
HCBD
HCBD
HCBD
HCBD
PCBG
HCBD
HCBD
Chromosome aberrations
Chromosome aberrations
Sister chromatid
exchange
Unscheduled DNA
synthesis
Unscheduled DNA
synthesis
Morphological
transformation
Unscheduled DNA
synthesis
+/- rat liver S9
+/- rat liver S9
+/- rat liver S9
+/- activation
Addition of acivicin or
aminooxyacetic acid
gave negative results
+/- activation
-
-
+
+
+
+
-
Galloway etal. (1987)
German (1988)
Galloway etal. (1987)
Schiffmann et al. (1984)
Vamvakas etal. (1989)
Schiffmann et al. (1984)
Stottetal. (1981)
In Vivo Assays
Mouse; bone
marrow cells
Mouse; bone
marrow cells
Rat; bone marrow
cells
HCBD
HCBD
HCBD
Chromosome aberrations
Chromosome aberrations
Chromosomal aberration
Inhalation, 4h
Oral gavage
Dietary doses of up to
20 mg/kg-day for 148
days
+
+
-
German (1988)
German (1988)
Schwetz etal. (1977)
HCBD February 2003
7-20
-------
Table 7-4 (continued)
Species/Strain/
Cell Type
Rat; bone marrow
cells
Rat
Rat kidney cells
Rat kidney cells
Mouse
Drosophila
melanogaster
Drosophila
melanogaster
Compound
HCBD
HCBD
HCBD
HCBD
HCBD
HCBD
HCBD
End Point
Chromosomal aberrations
Dominant lethality
DNA alkylation
DNA repair
DNA binding
Gene mutation (sex-
linked recessive lethal)
Sex-linked lethals
Comments
10 or 50 ppm, 7 hrs/day,
1 or 5 days
10 or 50 ppm, 7 hrs/day,
5 days
20 mg/kg by gavage
20 mg/kg by gavage
Single dose, 30 mg/kg
Feeding or injection
Results3
-
-
+
+
+b
-
-
Reference
NIOSH (1981)
NIOSH (1981)
Stottetal. (1981)
Stottetal. (1981)
Schrenk and Dekant (1989)
NIOSH (1981)
Woodruff etal. (1985)
Source: modified from ATSDR (1994) and WHO (1994).
a + = > twice the background rate or, in the case of bacterial studies, a reproducible dose-related increase in the number of revertants per plate;
nd = not determined;
b Binding was predominately to mitochondrial DNA
- = negative;
HCBD February 2003
7-21
-------
In Vivo Test Systems
Results of in vivo HCBD genotoxicity tests are summarized in the lower portion of Table 7-
4. Both negative (NIOSH, 1981; Schwetz et al., 1977) and positive (German, 1988) results have
been reported for chromosome aberration assays conducted in HCBD-treated mice or rats. Negative
findings have been reported in a dominant lethal assay in rats (NIOSH, 1981), and in the Drosophila
melanogaster sex-linked recessive lethal mutation assay with exposure via either inj ection or feeding
(NIOSH, 1981; Woodruff et al., 1985). However, Stott et al. (1981) reported a small (1.25 to 1.54-
fold) increase in UDS activity and DNA alkylation (0.78 alkylation per 106 nucleotides) in kidney
cells from rats fed 20 mg/kg-day HCBD in the diet for 3 weeks, suggesting that HCBD exhibited
a minor degree of renal genotoxicity.
Schrenk and Dekant (1989) evaluated the covalent binding of HCBD metabolites to renal
and hepatic DNA in NMRI mice. A low level of covalent binding (covalent binding index (CBI) =
27) was observed with nuclear DNA (nDNA) isolated from the kidney, while covalent binding was
undetectable in nDNA isolated from liver. Significantly higher levels of covalent binding were
observed with mitochondrial DNA (mtDNA), with CBIs of 513 and 7,506 determined for liver and
kidney, respectively. Analysis of covalent binding to renal mtDNA identified three 14C-labeled
compounds that appeared to be DNA bases altered by HCBD metabolites.
7.3.2 Immunotoxicity
The immunological effects of HCBD have not been systematically evaluated in humans, and
there are currently no case reports that describe immunological abnormalities occurring in humans
exposed to HCBD.
Animal data on the immunological effects of HCBD are limited. In a 2-week oral exposure
study conducted by NTP (1991), depletion (atrophy) and necrosis of lymphoid tissues of the spleen,
thymus, and lymph nodes was observed in B6C3FJ mice administered lethal concentrations of 1,000
and 3,000 ppm of HCBD. However, the investigators noted that these lesions may have been
secondary to chemical-induced stress. Similar lesions were not observed in mice administered 19.2
mg/kg-day in a subsequent 13-week study conducted by NTP (1991).
In a 13-week gavage study, relative spleen weights were significantly increased in male rats
orally administered HCBD at 15.6 mg/kg-day and in females at 6.3 mg/kg-day and above (Harleman
and Seinen, 1979). Treatment-related lesions in lymphoid organs (thymus, lymph nodes, spleen)
have not been reported in terminal necropsy of mice or rats in other HCBD subchronic and chronic
oral exposure studies at doses up to 100 mg/kg-day (Harleman and Seinen, 1979; Kociba et al. 1977;
Schwetz et al., 1977). Immune function screening batteries in HCBD-treated animals has not been
evaluated.
Delayed hypersensitivity reaction was exhibited in guinea pigs to dermal HCBD application
(Gradiski et al., 1975).
HCBD February 2003 7-22
-------
7.3.3 Hormonal Disruption
No studies were identified that associate HCBD exposure with endocrine disruption.
7.3.4 Physiological or Mechanistic Studies
Many studies have been undertaken to investigate the mechanisms underlying the effects of
HCBD and, in particular, its toxicity in the proximal renal tubule. The mode of action for HCBD as
determined from these studies and implications for cancer assessment are discussed in Section 7.4.3
The proximal tubule-specific toxicity of HCBD is likely determined by two factors: 1) the
distribution of enzymes required for its bioactivation, and 2) the ability of this region to concentrate
precursors of the ultimate toxic species (Dekantetal., 1990). The enzyme cysteine conjugate P-lyase
is believed to catalyze the conversion of HCBD-cysteine conjugates to a highly reactive thioketene
metabolite (Figure 6-1, Section 6.3). Multiple investigators have addressed the localization of P-
lyase and its relationship to nephrotoxicity. MacFarlane et al. (1989) demonstrated by
immunocytochemical technique that the region of highest cytosolic P-lyase activity in untreated rats
coincides with the site of HCBD-induced necrosis in the pars recta region of the proximal tubule.
However, Jones et al. (1988) and Kim et al. (1997) detected P-lyase in the entire proximal tubule.
Trevisan et al. (1998) detected histopathological changes and increased levels of P-lyase activity in
the urine following treatment of rats with S3 and SrS2 specific nephrotoxicants, which were cited
as evidence for distribution of the enzyme along the entire length of the proximal tubule. These data
suggest that additional factors may contribute to selective damage in the pars recta.
The ability of the proximal tubule to concentrate HCBD metabolites has been investigated
as a factor in renal toxicity. Nash et al. (1984) administered a single dose of radiolabeled HCBD and
observed that radioactivity was concentrated in the renal cortex shortly after dosing. Renal cells that
concentrated the radiolabeled compounds were subsequently observed to undergo necrosis. In
mammals, y-glutamyltranspeptidase, the enzyme that together with dipeptidase catalyzes the
conversion of the glutathione conjugate to the cysteine conjugate, is concentrated in the brush-border
membrane of the proximate tubular cells. The distribution of this enzyme may also contribute to an
increase in the concentration of cysteine conjugates in the proximal renal tubules.
The probenecid-sensitive organic anion transporter that is present on the basolateral side of
proximal tubule cells appears to play a role in the accumulation of HCBD metabolites (Dekant
1996). Probenecid is a competitive inhibitor of organic anion transport, and is reported to be without
effect on energy metabolism, transport carrier synthesis, or uptake of other substances actively
transported by the kidney (Lock and Ishmael, 1985; Dekant, 1996). Haloalkene-derived
mercapturates have the highest affinity for the organic anion transporter, but glutathione and
cysteine S-conjugates with lipophilic substituents on sulfur are also substrates for transport (Dekant,
1996).
The effect of probenecid on development of HCBD-induced renal toxicity has been
investigated in in vivo and in vitro studies. Lock and Ishmael (1985) administered a single
intraperitoneal dose of 16or64|imol/kg14C-radiolabeled7V-acetyl-pentachlorobutadienyl-L-cysteine
to female Alpk/AP rats and observed acute renal necrosis within four hours. Prior administration of
HCBD February 2003 7-23
-------
up to 500 |imol/kg probenecid reduced renal cortical concentrations of radioactivity and provided
protection against nephrotoxicity in a dose-dependent manner as assessed by plasma urea
concentration and renal histopathology. Pretreatment with probenecid also reduced or prevented the
toxic effects of intraperitoneally injected HCBD and its glutathione and cysteine conjugates. Thus,
the selective toxicity to the pars recta in rats is thought to result in part from transport of HCBD
metabolites into cells of this region via a probenecid-sensitive transport system.
Bach et al. (1986) confirmed antagonism of probenecid in HCBD metabolite-induced toxicity
in freshly isolated rat proximal tubule fragments. Incorporation of 3H-proline into acid precipitable
protein was utilized as an indicator of synthetic capacity of the tubular fragment. Addition of 2 mM
TV-acetyl-pentachlorobutadienyl-L-cysteine to the incubation medium reduced 3H-proline
incorporation to 34% of the control value. The inclusion of 400 |j,M probenecid in the incubation
medium almost completely restored (to 95%) 3H-proline incorporation.
Multiple studies suggest that renal cortical mitochondria are a primary subcellular target for
HCBD toxicity. Jones et al. (1986) investigated the toxic effects of pentachlorobutadienyl-
glutathione (PCBG) in isolated rat renal epithelial cells. Exposure to PCBG decreased cell viability
and reduced the concentration of intracellular thiols. Other PCBG-related effects included depletion
of Ca2+from the mitochondrial compartment, an elevation of cytosolic Ca2+ concentration, inhibition
of respiration, and decreased levels of ATP. Prevention of PCBG bioactivation by inhibition of y-
glutamyl transpeptidase or P-lyase provided complete protection against cytotoxicity. The authors
hypothesized that PCBG-induced renal cell injury results from selective effects on mitochondrial
function, including inhibition of respiration, depression of ATP synthesis, and release of
mitochondrial calcium (II) ions.
Wallin et al. (1987) studied S-pentachlorobutadienyl-L-cysteine (PCBC) toxicity in
mitochondria isolated from the rat kidney cortex. Respiring mitochondria exposed to PCBC showed
a dose-dependent loss of ability to retain calcium. This effect was associated with a collapse of
mitochondrial membrane potential. A slow nonenzymatic depletion of mitochondrial glutathione
was also observed. Preincubation with aminooxyacetic acid, an inhibitor of P-lyase, effectively
counteracted the loss of glutathione, suggesting that an interaction of the reactive thioketene with
the mitochondrial inner membrane was responsible for the observed effects.
Schnellmann et al. (1987) investigated the mechanism of PCBC-induced toxicity in renal
proximal tubules isolated from New Zealand rabbits. Suspensions of isolated tubules were exposed
to concentrations of 20 to 500 |j,M PCBC in the presence or absence of chemicals which targeted
the activity of ion-channel ATPases (nystatin, ouabain), Cytochrome c (ascorbate,
tetramethylphenylenediamine), cytochrome oxidase (carbonyl cyanide p-
trifluoromethoxyphenylhydrazone) and/or ATP synthase (atractyloside or oligomycin). Fifteen
minutes of exposure to 200 |j,M PCBC caused an increase in basal and ouabain-insensitive
respiration but not ny statin-treated respiration. However, sixty minutes of PCBC exposure inhibited
basal, nystatin-stimulated and ouabain insensitive respiration, and resulted in a 79% decrease in
glutathione concentration. In addition, at 60 minutes, an 11% decrease in lactate dehydrogenase
retention was observed, suggesting that cell viability was decreased over time as a result of
treatment. The changes in respiration observed at 60 minutes appeared to result from gross
mitochondrial damage characterized by inhibition of state 3 (aerobic) respiration, depression of
HCBD February 2003 7-24
-------
cytochrome c/cytochrome oxidase activity, and inhibition of electron transport. The results of these
studies suggest that alterations in mitochondrial function are an early event in PCBC-mediated
toxicity.
A similar pattern of events was observed by Groves etal. (1991). These workers investigated
the relationship between uptake and covalent binding of the HCBD metabolite
pentachlorobutadienyl-L-cysteine (PCBC) in rabbit renal proximal tubules, renal membrane vesicles,
and isolated renal cortical mitochondria. Their findings confirmed the PCBC-induced pattern of
mitochondrial dysfunction previously observed by Schnellmann et al. (1987) in rabbit proximal
tubule suspensions. Furthermore, Groves et al. (1991) demonstrated the rapid accumulation of 35S-
PCBC in renal proximal tubule cells and its metabolism to a reactive intermediate that bound to
tubular protein. An estimated 70 to 90% of the intracellular radioactivity was bound to protein.
Mitochondria isolated from renal proximal tubules also metabolized 35S-PCBC to a reactive
intermediate that bound to mitochondrial protein, consistent with the mitochondrion being a critical
subcellular target for HCBD-induced toxicity. Addition of the P-lyase inhibitor aminooxyacetic acid
(AOAA) reduced covalent binding to tubular proteins, and blocked the toxic effects of PCBC on
isolated mitochondria. However, AOAA decreased but did not prevent the toxic effects PCBC on
respiration and cellular ATP levels induced by PCBC exposure.
Additional studies have investigated interactions between the reactive intermediate generated
by metabolism of PCBC and cellular macromolecules. Lock and Schnellmann (1990) examined the
ability of reactive thiols formed by the action of P-lyase on cysteine conjugates of several
haloalkenes, including HCBD, to inhibit renal enzymes. The activities of glutathione reductase (a
cytosolic enzyme) and lipoyl dehydrogenase (a mitochondrial enzyme) were assayed for this
purpose. Administration of 200 mg/kg HCBD to male rats by intraperitoneal injection resulted in
inhibition of both enzymes. The authors suggested that such inhibition is a general outcome of
PCBC exposure, and is likely to occur with a diverse range of renal enzymes.
Schrenk and Dekant (1989) investigated covalent binding of 14C-labeledHCBD metabolites
to mouse DNA after a single oral dose of 30 mg/kg 14C-HCBD. HCBD metabolites bound
extensively to mitochondrial DNA. In contrast, little binding to nuclear DNA was observed. The
study authors suggested that proximity to high P-lyase concentration in the mitochondrial membrane
and the absence of associated histones make mitochondrial DNA a more vulnerable target for
reactive HCBD metabolites.
As noted above (Section 6.3), the cysteine derivative of HCBD is a substrate for cysteine
conjugate P-lyase. The activity of P-lyase leads to formation of an enethiol intermediate which is
rapidly converted to thioketene, a potent acylating agent (Dekant et al., 1990). In rats, the enethiol
intermediate may be detoxified by methylation to form pentachlorobutadienyl-methylthioether.
Morel et al. (1999) investigated the role of ^-adenosyl methionine (SAM)-dependent thiol
methylation in prevention of HCBD-induced nephrotoxicity in male Swiss OF 1 mice. The mice were
treated with a single intraperitoneal dose of periodate-oxidized adenosine (ADOX) prior to
administration of a single intraperitoneal dose of 80 or 100 mgHCBD/kg. Pretreatment with ADOX
increased the level of SAM in the liver and kidney approximately four-fold, but did not modify the
nephrotoxicity of HCBD as determined by histopathological evaluation of renal proximal tubules.
HCBD February 2003 7-25
-------
This result was interpreted by the study authors as evidence that SAM-dependent thiol methylation
does not play a role in detoxification of HCBD-derived enethiol in mice.
Chemically-induced a2(J-globulin nephropathy represents a potential alternative mechanism
for HCBD toxicity in rats. Since a2(J-globulin synthesis is androgen-dependent in the liver, this form
of nephropathy occurs exclusively in male rats, and is characterized by the accumulation of hyaline
droplets in proximal tubule cells. Binding of the chemical to oc2(J-globulin is a prerequisite for
development of nephrotoxicity. Because HCBD-induced renal toxicity occurs in both male and
female rats, it is evident that oc2(J-globulin nephropathy is not required for nephrotoxicity. However,
there is limited evidence to suggest that the a2(J-globulin mechanism may contribute to HCBD-
induced nephrotoxicity observed in male rats. Birner et al. (1995) observed that unmetabolized
HCBD was excreted in the urine of male, but not female, Wistar rats following exposure to a single
gavage dose of 14C-HCBD in corn oil. The study authors also noted more pronounced necrotic
changes in the proximal tubules of male rats when examined 48 hours after treatment. Slight liver
damage was observed only in male rats. In a subsequent experiment in the same laboratory, Pahler
et al. (1997) orally administered 200 mg/kg 14C-HCBD in corn oil to Sprague-Dawley (SD) and NCI
Black-Reiter rats (NBR), an a2(J-globulin-deficient strain 14C-HCBD was present only in the urine
of male SD rats, but not NBR rats. The study authors determined that the excreted HCBD detected
in the urine of male SD rats was associated with its binding to oc2(J-globulin. Histopathological
examination 48 hours after treatment revealed the formulation of hyaline droplets indicative of a2(J-
globulin accumulation in renal epithelial cells of male SD rats, but not in male NBR rats. In addition,
microscopic examination confirmed the occurrence of more extensive nephropathy in male than in
female animals as previously observed in Wistar rats.
Saitoetal. (1996) established that dose-dependent levels of kidney -type a2(J-globulin (aG-K)
in the urine are a reliable predictor of oc2(J-globulin accumulation in the kidney. These investigators
subsequently administered 100 mg/kg-day HCBD to adult male Sprague-Dawley rats for five
consecutive days. No increase in urinary ocG-K was detected following exposure, suggesting that
HCBD treatment did not induce a marked accumulation of a2(J-globulin. However, histopathological
examination revealed some epithelial cells showing hyaline droplet-related degeneration. The size
of the hyaline droplets formed following HCBD-treatment were generally smaller than those
observed after treatment with the well-characterized oc2(J-globulin nephropathy-inducing agent d-
limonene.
No significant increase in oc2(J-globulin was observed in kidney cytosol prepared from Fischer
344/N rats treated with a single oral dose of 200 mg HCBD/kg when assayed by Western blot and
capillary electrophoresis (Pahler et al., 1999).
oc2(J-globulin accumulation is unique to male rats, and when in excess is associated with
nephropathy, cytotoxicity, cellular proliferation, and consequent renal tumors. HCBD does not cause
an accumulation of oc2(J-globulin in the male rats (Saito et al., 1996; Pahler et al., 1999). Moreover,
HCBD clearly causes kidney damage in both male and female rats and in mice (Kociba et al., 1977;
NTP, 1991). With respect to kidney tumors, Kociba (1977) shows that kidney tumors are found in
both HCBD-treated male and female rats. Based upon current EPA guidelines (U.S. EPA, 1991e),
oc2(J -globulin accumulation does not appear to be a mode of action of HCBD-induced kidney toxicity
in rats or relevant to human risk.
HCBD February 2003 7-26
-------
7.3.5 Structure-Activity Relationship
HCBD appears to share a common target tissue with several structurally-related haloalkenes,
including perfluoropropene, trichloroethene, tetrachloroethene, and trichlorotrifluoropropene. All
of these chemicals demonstrate dose-dependent toxicity in the proximal tubule. Trichloroethene,
tetrachloroethene, tetrafluoroethylene, and HCBD have all been found to induce neoplasia of the
proximal tubule in rats. The common basis for toxicity may be bioactivation of these compounds
by a multistep pathway which is initiated by conjugation with glutathione, resulting in the formation
of a glutathione ^-conjugate. Metabolism to the corresponding cysteine ^-conjugates, and
subsequent degradation by renal cysteine conjugate P-lyase, yields reactive electrophiles that are
believed to be ultimately responsible for renal toxicity. These electrophiles alkylate mitochondrial
macromolecules, resulting in cellular energy deficit, loss of membrane potential, and disruption of
calcium homeostasis.
When haloalkenes are considered as a group, the extent of conjugation is much higher with
liver microsomes than with liver cytosol, in contrast to results observed with most other substrates.
This effect is attributed to the preferential distribution of the highly lipophilic haloalkenes into lipid
membranes, thus providing high substrate concentrations for membrane-bound glutathione S-
transferase (Dekant et al., 1990). In vitro studies suggest that rates of haloalkene conjugation
correlate well with the chemical reactivity of the individual compounds (Dekant et al., 1990). For
example, substitution with chlorine results in stabilization of a u-bond. Chloroalkenes are thus
reported to be more resistant to metabolism by glutathione conjugation than are fluoroalkenes.
Investigations have compared the toxicity of structurally-related haloalkene conjugates.
Lock and Schnellmann (1990) investigated the effect of HCBD and other haloalkene cysteine
conjugates on renal glutathione reductase and lipoyl dehydrogenase activity, and concluded that
inhibition of these enzymes by the reactive thiols formed by P-lyase cleavage of haloalkene cysteine
conjugates represented a general mechanism of toxicity.
Green and Odum (1985) investigated the nephrotoxicity and mutagenicity of the cysteine
conjugates of halogenated alkenes in rat kidney slices. Compounds investigated included the
chloroalkenes HCBD, trichloroethylene and perchloroethylene, and the fluoroalkenes
hexafluoropropene (HFP) and tetrafluoroethylene (TFE). All of these conjugates had a marked effect
on the uptake of both the organic anion /7-aminohippuric acid (PAH) and the cation
tetraethylammonium bromide (TEA) into rat kidney slices. This observation was considered to be
consistent with the known nephrotoxicity of HCBD, TFE and HFP in vivo. Each of the conjugates
was metabolized by rat kidney slices and by semi-purified rat kidney P-lyase to pyruvate, ammonia,
and an unidentified reactive metabolite. Although all of the conjugates were activated by P-lyase and
had a similar effect on ion transport, their mutagenicity differed. The conjugates of HCBD,
trichloroethylene and perchloroethylene were mutagenic in the Ames bacterial mutation assay when
activated by rat kidney S9 fraction. In contrast, the conjugates of TFE and HFP were not mutagenic
either in the presence or absence of rat kidney S9 fraction.
Birner et al. (1997) compared the nephrotoxicity of cysteine ^-conjugates derived from
trichloroethene, tetrachloroethene, and HCBD. Male and female rats received identical intravenous
HCBD February 2003 7-27
-------
doses of S-O^-dichlorovinyO-L-cysteine (1,2-DCVC), ^-(l^-dichlorovinyO-L-cysteine (2,2-
DCVC), S-(l,2,2-trichlorovinyl)-L-cysteine (TCVC), or S-(l,2,3,4,4-pentacMorobutadienyl)-L-
cysteine (PCBC). Assessment of the relative nephrotoxic potency of the conjugates by
histopathological examination and excretion of y-glutamyltranspepidases in urine indicated a
decrease in the order TCVC > 1,2-DCVC > PCBC > 2,2-DCVC.
7.4 Hazard Characterization
7.4.1 Synthesis and Evaluation of Major Noncancer Effects
There are no unconfounded reports of human health effects following HCBD exposure by
any route. Oral exposure studies of HCBD toxicity in animals are summarized in Table 7-5. A
distinctive feature of HCBD toxicity in animals is its selective effect on the kidney, regardless of
the route of administration. Toxicity within the kidney is also selective, with damage restricted to
the proximal tubule. In rats, damage is further localized to the pars recta region of the proximal
tubule.
Subchronic and chronic studies in rodents present a clear picture of dose-related renal
damage. Progressive events overtime include changes in kidney weight, renal tubular degeneration,
necrosis and regeneration, hyperplasia, focal adenomatous proliferation, and tumor formation.
Tumor formation occurs exclusively in the kidney, and only at doses that cause extensive
cytotoxicity.
Evidence from metabolic enzyme inhibitor studies, cannulation experiments, and analysis
of urinary metabolites indicates that the nephrotoxicity of HCBD is dependent on a multistep
bioactivation mechanism involving both liver and kidney enzymes. The initial step in HCBD
metabolism is the glutathione-^-transferase mediated biosynthesis of a glutathione conjugate
(PCBG) in the liver. After elimination into the bile, PCBG undergoes subsequent metabolism to a
cysteine conjugate (PCBC) in the bile, gut or kidneys. PCBC may be acetylated by renal N-
acetyltransferases to form a7V-acetyl cysteine conjugates (7V-AcPCBC). Both PCBC andTV-AcPCBC
are concentrated in renal cells via an active transport system (Dekant, 1990). 7V-AcPCBC can be
excreted in the urine or de-acetylated to regenerate PCBC. PCBC is a substrate for p-lyase-
dependent activation to a highly reactive thioketene in the kidney. Covalent binding of this reactive
species to cellular macromolecules is believed to initiate the damage that ultimately results in renal
cell toxicity.
Potential molecular targets for binding of the reactive thioketene include enzymes,
membrane proteins, glutathione, phospholipids, and mitochondrial DNA. Localized damage to the
proximal tubule is believed to reflect high P-lyase concentration in this region. Evidence from
studies using the selective inhibitor probenecid suggests that accumulation of the cysteine and N-
acetyl cysteine conjugates via anion transport systems localized in this segment of the proximal
tubule may account for this selective pattern of toxicity.
In vitro studies suggest that cortical mitochondria are the critical subcellular target for
toxicity of the bioactivated sulfur conjugates of HCBD. Susceptibility of mitochondria to HCBD
toxicity is linked to high concentrations of P-lyase associated with mitochondrial membranes. The
HCBD February 2003 7-28
-------
Table 7-5. Summary of Principal HCBD Toxicity Studies.
Species/Strain
Number/
Sex/Dose
Route
Frequency/
Duration
Doses
(mg/kg-day)
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
Effect(s)
Reference
ACUTE EXPOSURE
Rat
Rat / Wistar
Rat / Sprague-
Dawley
6M
5M
4-5 M
G
G
G
Single Dose
24 hours
Single Dose
24 hours
Single Dose
24 hours
200
0, 10, 100,
200
0, 100, 200
10
..
200
100
100
Increased plasma urea;
increased proteins and
metabolites in urine
Limited focal necrosis of
the kidneys and other signs
of renal damage in urine
and blood
Histological kidney lesions
and impaired kidney
function
Nash et al.
(1984)
Jonker et al.
(1993a)
Payan et al.
(1993)
SHORT-TERM EXPOSURE
Rat / Sprague-
Dawley
Rat / Wistar-
derived
Rat / Sprague-
Dawley
4F
6M
6F
5M
F
F
G
Daily
30 days
Daily
14 days
Daily
3 weeks
0, 1, 3, 10, 30,
65, 100
0, 4.6, 14,
35.3
0, 0.2, 20
3
0.2
10
(reduced body wt.
gain; increased
hemoglobin
concentration)
4.6
20
Increased relative kidney
weight; renal tubular
degeneration, necrosis and
regeneration; increased
hemoglobin concentration
reduced body weight gain
Degeneration of renal
tubular epithelial cells,
localized to pars recta
Histopathological
indications of renal cortical
damage; decreased body
weight gain; increased
relative kidney weight
Kociba et al.
(1971)
Harleman
and Seinen
(1979)
Stott et al.
(1981)
HCBD February 2003
7-29
-------
Table 7-5 (continued)
Species/Strain
Mouse /
B6C3F!
Rat / Wistar
Rat / Wistar
Number/
Sex/Dose
5M
5F
5M
5F
12 M
Route
F
F
F
Frequency/
Duration
Daily
15 days
Daily
4 weeks
Daily
3 weeks
Doses
(mg/kg-day)
M: 0, 3, 12, 40
F: 0, 5, 16, 49
0, 2.25, 8, 28
0, 7.2, 36, 180
NOAEL
(mg/kg-day)
-
2.25
7.2
LOAEL
(mg/kg-day)
3
5
8
36 (body weight)
180 (kidney lesions)
Effect(s)
Renal necrosis and cellular
regeneration
Renal tubular cytomegaly;
decreased plasma creatine;
decreased body weight,
decreased liver and adrenal
weight
Lower mean body weight
(15% decrease); extensive
regeneration at 180 but not
36 mg/kg-day
Reference
NTP (1991)
Yang et al.
(1989)
Jonker et al.
(1993b)
Nakagawa et
al. (1998)
SUBCHRONIC EXPOSURE
Rat / Wistar-
derived
Mouse /
B6C3FJ
10 M
10 F
10 M
10 F
G
F
Daily
13 weeks
Daily
13 weeks
0, 0.4, 1.0,
2.5,6.3, 15.6
M: 0,0.1,0.4,
1.5,4.9, 16.8
F: 0,0.2,0.5,
1.8,4.5, 19.2
2.5 (M)
1.0 (F)
1.5 (M)
0.2? (F)
6.3 (M)
2.5 (F)
4.9 (M)
0.5 (F)
Proximal tubular
degeneration
Renal tubular regeneration
Harleman
and Seinen
(1979)
NTP (1991)
HCBD February 2003
7-30
-------
Table 7-5 (continued)
Species/Strain
Rat / Sprague-
Dawley
Rat / Wistar
Number/
Sex/Dose
Gross
Pathology:
10-12 M,
20-24 F
(HCBD)17
M, 34 F
(control)
Histopatho
logy:
5M
5F
21 M
Route
F
F
Frequency/
Duration
Daily
148 days
Daily
30 weeks
Doses
(mg/kg-day)
0.2, 2.0, 20
90
NOAEL
(mg/kg-day)
0.2
LOAEL
(mg/kg-day)
2
90
Effect(s)
Renal tubular hyperplasia
Decreased final weight and
increased relative kidney
weight
Reference
Schwetz et
al. (1977)
Nakagawa et
al. (1998)
CHRONIC EXPOSURE
Rat / Sprague-
Dawley
39M, 40F
F
Daily
22 months (M)
24 months (F)
0, 0.2, 2, 20
0.2
2
Increased urinary
coproporphyrin excretion
(females); increased kidney
weight; renal tubular
epithelial hyperplasia
Kociba et al.
(1977)
HCBD February 2003
7-31
-------
Table 7-5 (continued)
Species/Strain
Number/
Sex/Dose
Route
Frequency/
Duration
Doses
(mg/kg-day)
NOAEL
(mg/kg-day)
LOAEL
(mg/kg-day)
Effect(s)
Reference
DEVELOPMENTAL AND REPRODUCTIVE STUDIES
Rat / Sprague-
Dawley
Rat
Rat
10-12 M,
20-24 F
(HCBD)
17M, 34 F
(control)
6F
F
F
F
NS
Daily
148 days (90
days prior to
mating through
postnatal day
21)
Daily
3 weeks prior
to mating; 3
weeks during
mating;
throughout
gestation and
lactation
NS
0, 0.2, 2, 20
0, 15, 150
8.1
2
20
15
8.1
Slightly decreased neonatal
weight
Maternal toxicity (renal,
neurological). Reduced pup
weights on days 0, 10, and
20
Reduced body weight;
shorter crown-rump length;
ultrastructural chances in
neurocytes; increased levels
of free radicals in brain and
spinal cord
Schwetz et
al. (1977)
Harleman
and Seinen
(1979)
Badaeva
(1983)
Abbreviations: Sex: M = male
F = female
Route: G = gavage
F = feed study
NS = Not specified
HCBD February 2003
7-32
-------
reactive metabolite formed by p-lyase cleavage of sulfur conjugates is thought to interact with
components of the inner mitochondrial membrane. Disruption of respiration and uncoupling of
oxidative phosphorylation leads to a marked reduction of ATP levels in susceptible kidney cells, and
ultimately necrosis.
The mechanism described above is believed to contribute to the renal damage observed in
both male and female rats. However, additional mechanisms may contribute to nephrotoxicity of
HCBD in male rats. Current evidence suggests that at least two discrete male-specific pathways may
participate in the more pronounced necrotic changes observed in the renal tubules of male rats in
some studies. Formation of hyaline droplets indicative of oc2(J-globulin accumulation has been
observed in the kidney of HCBD-treated male rats. The significance of this finding for HCBD-
induced nephrotoxicity remains to be determined. A second potential mechanism for male specific
toxicity involves the cytochrome P450 3 A-mediated formation of anTV-acetylated cysteine conjugate
sulfoxide.
Other noncancer effects associated with HCBD exposure in animals include developmental
effects and neurotoxicity. Reproductive effects were observed only at maternal toxic dose. In one
study, female Wistar rats were administered a diet containing 0, 15 or 150 mg/kg-day) HCBD for
3 weeks prior to mating, 3 weeks during mating and throughout gestation and lactation. Maternal
toxicity was evident in treated groups. No conceptions occurred for the high dose group, the ovaries
showed little follicular activity, and no uterine implantation sites were observed. At 15 mg/kg-day,
pups exhibited lower birth weights and reduced growth compared to controls (Harleman and Seinen,
1979). In a Russian study cited through secondary sources, pregnant rats administered 8.1 mg/kg-
day of HCBD during gestation gave birth to pups with lower body weights and shorter crown-rump
lengths (Badaeva, 1983).
Harleman and Seinen (1979) observed ataxia, incoordination, weakness of the hind legs, and
unsteady gait in conjunction with demyelination and fragmentation of femoral nerve fibers in female
rats consuming dietary dose of 150 mg/kg-day HCBD for 10 to 18 weeks. No neurotoxic effects
were reported for rats consuming 15 mg/kg-day. In a Russian study cited through secondary sources,
daily oral administration of 8.1 mg/kg-day HCBD to pregnant rats throughout gestation resulted in
histopathological changes in nerve cells and myelin fibers of the spinal cord in treated dams and
their offspring (Badaeva et al., 1985).
The mode of neurotoxic action has not been studied. Toxicokinetic studies in animals
following oral administration demonstrated that HCBD and its metabolites distributed to the brain
and adipose tissues in addition to the kidney and the liver (Reichert, 1983; Reichert et al., 1985;
Dekant et al., 1988a). Thus, reported toxicity at targets other than the kidney may be related to the
distribution of HCBD and/or its metabolites to these targets and subsequent covalent binding of the
reactive metabolites to cellular macromolecules.
An important issue in the evaluation of the hazard posed by HCBD concerns the applicability
of mechanistic data obtained in rodent studies to humans. Limited data from in vitro studies with
human renal cytosol and cultured human proximal tubule cells suggest that humans have the ability
to form HCBD glutathione conjugates and to metabolize HCBD cysteine conjugates to a toxic
metabolite.
HCBD February 2003 7-33
-------
7.4.2 Synthesis and Evaluation of Carcinogenic Effects
No studies of the potential for HCBD carcinogenicity in humans have been reported. In
animals, one lifetime exposure carcinogenicity study has been performed. Kociba et al. (1977)
observed increased incidence of renal tumor formation in male and female rats following lifetime
exposure to HCBD in the diet. Neoplastic changes occurred only at the highest dose, which
exceeded the maximum tolerated dose (MTD). There was increased mortality, significant weight
loss (greater than 10%), and severe renal toxicity. This pattern suggests that tumor formation may
be secondary to HCBD-induced cytotoxicity. This conclusion is supported by the study of
Nakagawa et al. (1998), who found increased cell proliferation and increased DNA synthesis in the
outer stripe and cortex of kidneys from HCBD-treated rats.
However, available data must be considered as too limited to support a conclusion with high
confidence. The widely-spaced doses (a 10-fold spacing between the highest and next lower dose)
in the Kociba et al. (1977) study, for example, did not provide the opportunity to confirm that
pronounced cytotoxicity is a prerequisite for tumorigenesis. Additional limitations in the database
include the absence of cell proliferation studies and limited in vivo data for mutagenesis. These
limitations prevent the use of cell-kinetic multistage (CKM) models in the analysis of cancer risk
(Bogen, 1989).
Results from mutagenicity studies with HCBD are mixed. In the presence of appropriate
metabolic activation conditions, HCBD and its metabolites are mutagenic in some, but not all,
studies. Thus, a genotoxic mode of action must be considered. The observation that high doses of
HCBD metabolites can bind to DNA weakly in vivo in mice (Schrenk and Dekant, 1989) strengthens
this conclusion somewhat. There is, however, still some question as to whether genotoxicity is the
primary mode of action at the lower concentrations at which nephrotoxic effects are first observed.
The weight of evidence suggests that both genotoxic and nongenotoxic effects contribute, possibly
at different dose levels, and that the current database is insufficient to disentangle the relative
contributions of these effects.
7.4.3 Mode of Action and Implications in Cancer Assessment
Both sustained cytotoxic damage and irreversible DNA binding have been proposed as
events in HCBD renal carcinogenesis in rodents (Stott et al., 1981)An. A lifetime oral study in rats
showed kidney tumors at a very high dose that exceeded the MTD, suggesting that HCBD-induced
cytotoxicity may lead to tumor formation (see Section 7.4.2). Studies in rats and mice indicate that
kidney is the target organ. Progressive toxicological changes are observed in kidney over time:
decreased and increased kidney weight, increased excretion of coporphyrin (kidney dysfunction),
renal tubular degeneration, necrosis and regeneration, hyperplasia, focal adenomatous proliferation,
and finally tumor formation.
On the other hand, in the presence of metabolic activation, HCBD and its reactive
metabolites are mutagenic in some (Simmon, 1977;Reichertetal., 1984; Reichert and Schutz, 1986;
Wild et al., 1986), but not all, studies (See Section 7.3.1). Those studies done either on kidney cells
(in vivo) or with kidney microsomes (in vitro) appear to represent most of the mutagenic response
to HCBD. Thus, a mutagenic mode of action cannot be ruled out (Dekant et al., 1990; Lock, 1994).
HCBD February 2003 7-34
-------
The hypothesis that both cytotoxicity and mutagenic mode of action may be operating is
consistent with the findings that the adverse effects of HCBD are dependent on a multistep pathway
of bioactivation. The ultimate step in this pathway is a P-lyase-mediated degradation of a HCBD
metabolite that generates a highly reactive thioketene in proximal tubule cells. Covalent binding of
this thioketene to DNA, proteins and other macromolecules is considered to be the mechanism
responsible for the observed cytotoxic and mutagenic effects of HCBD and its metabolites.
Restriction of these effects to the proximal tubule most likely reflects both uptake processes that
concentrate the cysteine conjugate substrate in epithelial cells, and localization of y
glutamyltranspeptidase and P-lyase activity to this region of the kidney.
In vitro studies (Schnellman et al., 1987; Groves et al., 1991; Jones et al., 1986; Wallin et
al., 1987) indicate that mitochondria in renal tubular epithelial cells are the major target for HCBD
metabolite-induced toxicity. The reactive metabolite formed by P-lyase cleavage of cysteine
conjugate is thought to interact with components of the mitochondrial inner membrane. The initial
effect is an uncoupling of oxidative phosphorylation and prevention of generation of ATP. The
decrease in renal tubular ATP secondary to mitochondrial dysfunction in turn limits ATP dependent
active transport in the tubules, inhibiting reabsorption processes (Jaffe et al., 1983). Later effect
involves gross mitochondrial damage characterized by inhibition of cytochrome c-cytochrome
oxidase activity, and inhibited electron transport. This sequence of events in the renal proximal
tubules ultimately leads to cell death. Other studies indicate that the reactive species generated by
P-lyase-mediated degradation of HCBD metabolites interact directly with mitochondrial DNA
(mtDNA) from mouse kidney (Schrenk and Dekant, 1989). Renal mtDNA may be the preferential
target due to the high concentration of P-lyase in the mitochondrial membrane, the lack of protective
histones associated with mitochondrial DNA (Borst & Grivell, 1978), and an inadequate repair
function (Mansouri et al., 1997). Mutations in the mtDNA can lead to a respiratory chain deficiency
and cell dysfunction when the percentage of the mutants reach a certain level (Schapira, 1999).
Three important aspects of mitochondrial oxidative phosphorylation involved in the
pathogenesis of mitochondrial dysfunction are: generation of cellular energy in the form of ATP;
generation of reactive oxygen species (ROS); and regulation of apoptosis or programmed cell death
(Wallace, 1999). The process of oxidative phosphorylation produces significant amounts of ROS
which are toxic byproducts of respiration. Chronic exposure to ROS can result in oxidative damage
to mitochondrial and cellular proteins, and mutations in the mtDNA. Because mtDNA codes for 22
transfer RNAs (tRNA) and 2 ribosomal RNAs (rRNA) for synthesis of important mitochrondrial
proteins involved in the oxidative phosphorylation, functional mtDNA is critical to the normal
function of a cell. Mutations in mtDNA may lead to overexposure to ROS and decreased energy
production. Apoptosis is initiated when the mitochondrial permeability transition pore (mtPTP) in
the inner membrane opens and cell death-promoting factors such as the caspases are released
(Wallace, 1999). Opening of the mtPTP and the accompanying cell death can be initiated by the
mitochondrion's excessive uptake of Ca2+, increased exposure to ROS, or decline in energetic
capacity. Therefore, a marked reduction in mitochondrial energy production and a chronic increase
in oxidative stress could activate the mtPTP and initiate apoptosis.
Numerous mtDNA mutations have been associated with human mitochondrial disease.
Mitochondrial disease is a disruption of the proper function of the mitochondria, resulting in a
variety of clinical manifestation. This disruption can include an inhibition of the electron transport
HCBD February 2003 7-35
-------
chain, a disruption of oxidative phosphorylation and an increase in the production of reactive oxygen
species. MtDNA mutations could contribute to neoplastic transformation by changing cellular
energy capacities, mitochondrial oxidative stress, and/or modulating apoptosis (Wallace, 1999).
Thus, it may be postulated that mutations of renal mtDNA induced by HCBD may result in
reduction in energy production, increase in oxidative stress, and initiation of apotosis, leading to
tumor formation.
Mitochondrial dysfunction may also result from interaction of highly reactive HCBD
metabolites with components of the mitochondrial inner membrane, such as enzymes related to cell
function. Subsequent energy depletion may trigger the renal cytotoxicity that is the putative
mechanism forHCBD-mediated carcinogenesis. Thus, HCBD induced cytotoxicity andtumorigensis
may be ultimately the consequence of mitochondrial dysfunction resulting from exposure.
Recent evidence suggests that HCBD-induced a2(J-globulin accumulation contributes to renal
injury in male rats. However, renal tubular necrosis and renal tubular tumors were observed in both
male and female rats following HCBD exposure (Kociba et al., 1977), and renal necrosis and
regeneration were also observed in male and female mice (NTP, 1991). Therefore, a2(J-globulin
accumulation cannot be the sole mechanism for HCBD-induced carcinogenesis.
7.4.4 Weight of Evidence Evaluation for Carcinogenicity
No human carcinogenicity data are available for HCBD.
A single lifetime study of HCBD carcinogenicity in rats (Kociba et al., 1977) is available for
evaluation. This study revealed statistically significant increases in the incidence of tumors in male
and female rats following oral HCBD exposure. Although human carcinogenicity data are
unavailable, evidence exists that the metabolic enzymes responsible for conversion of HCBD to the
reactive and toxic thioketene occur in humans, albeit at levels lower than that in the rat (see Section
6.3). In accordance with EPA's 1986 Guidelines for Carcinogen Risk Assessment (U. S. EPA, 1986),
HCBD is best classified as Group C, possible human carcinogens, based on limited evidence of
carcinogenicity in one animal study, and no data in humans. Based on the proposed guidelines for
Carcinogen Risk Assessment (U.S. EPA, 19991996a), HCBD is classified as "likely to be
carcinogenic to humans by the oral route of exposure, but whose carcinogenic potential by the
inhalation and dermal routes of exposure cannot be determined because there are inadequate data
to perform an assessment'''. This descriptor is considered appropriate when there are no or
inadequate data in humans, but the combined experimental evidence demonstrates the production
or anticipated production of tumors in animals by modes of action that are relevant or assumed to
be relevant to humans, tempered by the lack of adequate inhalation or dermal studies. Mechanistic
studies performed in vitro and in vivo suggest that either genotoxic or non-genotoxic modes of action
may underlie or contribute to the carcinogenic potential of HCBD. However, the strongly non-linear
dose-response in the Kociba et al. (1977) study cannot be ignored. At present, the mode-of-action
information still lacks identification of the sequence of key events and a quantitative description of
the doses at which those key events begin to occur. In such cases, EPA's proposed cancer guideline
revisions (U. S.EPA, 1999) support consideration of both linear and nonlinear extrapolation to lower
doses. In the absence of adequate data to exclude a linear mechanism(s) of tumor formation, the
HCBD February 2003 7-36
-------
quantitative cancer risk assessment of HCBD should conservatively be conducted using the
linear-default model. Therefore, both linear and non-linear approaches are presented.
7.4.5 Sensitive Populations
Sensitive populations are those which experience more adverse effects at comparable levels
of exposure, or which experience adverse effects at lower exposure levels, than the general
population. The enhanced response of these sensitive 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.
Human populations that exhibit greater sensitivity to HCBD have not been identified.
However, it has been generally observed that existing nephropathy or age-related kidney
degeneration can increase the risk of renal injury or exacerbate nephrotoxicity in humans (WHO,
1991). Evidence that existing nephropathy increases sensitivity to HCBD toxicity has been obtained
in a study conducted in male Wistar rats (Kirby and Bach, 1995). Nephrosis was induced by
pretreatment with adriamycin (ADR), and rats were subsequently exposed to HCBD. Damage to
the proximal tubule was more severe and renal cortical repair capacity was decreased in ADR-
treated rats when compared to rats exposed to HCBD without prior ADR exposure. These results
suggest that individuals with existing kidney damage or the elderly may be potentially sensitive
populations for HCBD exposure.
Studies in animals showed that the young rats and mice experience acute effects at
significantly lower doses than do adults (Hook et al., 1983; Lock et al., 1984), suggesting that
infants may represent a potentially sensitive subpopulation for acute HCBD exposure, perhaps as
a result of immature organ systems. Additionally, female rodents were apparently more sensitive
than males to acute HCBD exposure (Kociba, 1977). The mechanism underlying this sensitivity is
not currently clear.
HCBD February 2003 7-37
-------
8.0 DOSE-RESPONSE ASSESSMENT
8.1 Dose-Response for Noncancer Effects
8.1.1 RfD Determination
The reference dose (RfD) for a chemical is "an estimate (with uncertainty spanning
approximately an order of magnitude) of a daily exposure to the human population (including
sensitive subgroups) that is likely to be without appreciable risk of deleterious effects over a
lifetime" (U.S. EPA, 1993). Data on the non-cancer effects of HCBD from chronic and subchronic
studies were used to estimate a RfD value using the benchmark dose (BMD2) approach (U. S. EPA,
1995).
Choice of Principal Study and Critical Effect
There are no reliable dose-response data for humans exposed to HCBD. The NTP (1991)
subchronic study on mice was chosen as the principal study, with the Kociba et al. (1977) and
Schwetz et al. (1977) studies on rats as supporting studies, due to their exposure durations and
sensitivity of endpoints observed. Hyperplasia and regeneration of the renal tubular cell epithelial
cells was selected as the critical effect, based upon its observation at the lowest doses. The RfD for
HCBD is derived from a BMDL of 0.1 mg/kg-day for renal tubular epithelial cell
hyperplasia/regeneration from the NTP study.
In the Kociba et al. (1977) lifetime oral exposure study of rats to HCBD, a NOAEL of 0.2
mg/kg-day and a LOAEL of 2 mg/kg-day were identified, based on an increase in renal tubular
epithelial cell hyperplasia/regeneration. In the Schwetz et al. (1977) 148-day oral exposure study
of rats to HCBD, aNOAEL of 0.2 mg/kg-day and a LOAEL of 2 mg/kg-day were identified, based
on an increase in the severity of renal tubular dilation, collapse, and atrophy. In the 13-week feeding
study by NTP (1991), the study authors identified a NOAEL of 1.5 mg/kg-day for male mice, and
did not identify a NOAEL for female mice because renal tubular regeneration occurred in 1 of 10
females in the lowest dose group (0.2 mg/kg-day). However, others (U.S. EPA, 1998a; WHO, 1994)
have concluded that the effect observed at 0.2 mg/kg-day is not statistically significant, and therefore
considered this dose to be the NOAEL.
In order to avoid the uncertainty in the LOAEL/NOAEL classification of the 0.2 mg/kg-day
dose in the NTP study, the most sensitive response found, a benchmark dose (BMD) analysis was
Note that the literature has used the terms BMD and BMDL in a confusing way (Crump, 1984,1995). The
EPA benchmark dose software (BMDS version 1.3.1) and EPA technical guidance on this subject (U.S.
EPA, 2000a), use the term "BMD" to refer to the central or maximum likelihood estimate (MLE) of the
dose that is expected to yield the BMR. "BMC" (benchmark concentration) is a term that is sometimes
used, as opposed to "BMD," to distinguish between inhalation and oral benchmarks. "BMDL" or "BMCL"
refer to the lower end of a one-sided confidence interval for that central estimate. "BMD" will be used to
refer to the entire process. The POD for low dose extrapolation on for setting the RfD/RfC will be the
BMDL or BMCL. To simplify further discussion in this document, we will use BMD or BMDL genetically
to mean oral or inhalation values, unless stated otherwise.
HCBD February 2003 8-1
-------
conducted on the NTP female mouse renal tubular regeneration response (Table 8-1). Using BMDS3,
version 1.3.1, the data was fit to all available dichotomous models (Table 8-2). The data was best
fit by a Weibull model (p = 1.00, AIC = 17), resulting in a BMD4 of 0.2 mg/kg-day and a BMDL
of 0.1 mg/kg-day (Figure 8-1). This dose was therefore taken as the point of departure for further
calculations.
Table 8-1. Incidence of Renal Tubular Regenerative Response in Mice Treated with HCBD
for 13 Weeks.
Test
Organism
Male Mice
Female Mice
Administered Dose in Feed
(ppm)
0
1
3
10
30
100
0
1
O
10
30
100
Approximate Daily
Dose
(mg/kg-day)
0
0.1
0.4
1.5
4.9
16.8
0
0.2
0.5
1.8
4.5
19.2
Renal Tubule
Regeneration
Incidence
0/10 (0%)
0/10 (0%)
0/10 (0%)
0/9 (0%)
10/10 (100%)
10/10 (100%)
0/10 (0%)
1/10 (10%)
9/10 (90%)
10/10 (100%)
10/10 (100%)*
10/10 (100%)*
* Not used in BMD analysis.
source: NTP (1991).
A copy of BMDS can be obtained from the Internet at http://www.epa.gov/ncea/bmds.htm.
A benchmark response (BMR) of 10% was used for all BMD analyses.
HCBD February 2003
8-2
-------
Table 8-2. Benchmark Dose Estimates from NTP (1991) Female Mouse Renal Tubular
Regeneration Response.
Model
Weibull
Gamma
Log-Probit
Log-Logistic
Probit
Logistic
Multistage (2)
Quantal -Quadrati c
Quantal-Linear
BMD
0.200
0.200
0.200
0.200
0.203
0.209
0.123
0.123
0.042
BMDL
0.099
0.111
0.126
0.123
0.112
0.119
0.056
0.094
0.026
Chi-
square
/7-value
1.00
1.00
1.00
1.00
0.99
0.97
0.68
0.68
0.14
AIC5
17
17
17
17
17
17
17
17
22
Akaike's Information Criterion (AIC) = -2L + 2p, where L is the log-likelihood at the maximum likelihood
estimates for the parameters, and p is the number of model degrees of freedom. This can be used to compare
models with different numbers of parameters using a similar fitting method (for example, least squares or a
binomial maximum likelihood). Although such methods are not exact, they can provide useful guidance in
model selection.
HCBD February 2003
8-3
-------
Figure 8-1. Benchmark Dose Estimate Using Weibull Model.
Weibull Model with 0.95 Confidence Level
0.8
T3
| 0.6
C
O
'C 0.4
CO
0.2
Weibull
BMD Lower Bound -T
BMD
0.5 1
Dose (mg/kg-day)
1.5
HCBD February 2003
8-4
-------
different areas of uncertainty. A composite uncertainty factor (UF) of 300 was used in the derivation
of the RfD. The composite UF included a factor of 10 to account for extrapolation from animals to
humans; a factor of 10 for protection of sensitive subpopulations; and a factor of 3 for database
deficiencies (lack of a 2-generation reproductive study6, and developmental toxicity studies in only
one species). Although some studies suggest that humans may have lower rates of formation of
putatively toxic metabolites than rodents, others suggest that the rate is the same (Section 6.3).
Calculation of RfD
Using the BMDL of 0.1 mg/kg-day from the NTP (1991) study, the RfD is derived as
follows:
RfD = (0.1 mg/kg-day) = 3 x 10"4 mg/kg-day
300
where:
0.1 mg/kg-day = BMDL, based on the histopathological effects in kidneys of mice
exposed to HCBD in the diet for up to 24 months (NTP, 1991).
300 = uncertainty factor. This is based on a factor of 10 to account for
extrapolation from animals to humans; a factor of 10 for protection
of potentially sensitive human subpopulations; and a factor of 3 for
database deficiencies (lack of a two-generation reproductive
study).
8.1.2. RfC Determination
RfC for HCBD is not derived. No subchronic or chronic inhalation exposure studies are
available for the determination of RfC.
8.2 Dose-Response for Cancer Effects
8.2.1 Choice of Study
As noted previously, only one lifetime oral carcinogenicity study of HCBD was located
(Kociba et al., 1977). In this study, Sprague-Dawley rats (40 animals/sex/dose group and 90
animals/sex in the control group) were dosed with 0, 0.2, 2 or 20 mg/kg-day HCBD via the diet for
22 months (males) or 24 months (females).
Neoplastic changes were found only at the highest dose, which exceeded the maximum
tolerated dose. There was a significant increase in mortality in males, a greater than 10% decrease
A single generation reproductive/developmental study misses important windows of vulnerability,
lactational exposure, and latent effects that only become evident as the Fl generation reaches maturity. That
is the rationale for the uncertainty factor.
HCBD February 2003 8-5
-------
in body weights for both sexes, and other severe renal toxicity effects were observed. The incidence
of renal tubular neoplasms was increased only in the high-dose group of both males and females,
as shown in Table 8-3.
Table 8-3. Incidence of Renal Tubular Neoplasms in Rats Treated with HCBD for 2
Years.
Test
Organism
Male
rats
Female rats
Administered
Dose
(mg/kg-day)
0
0.2
2.0
20
0
0.2
2.0
20
Human Equivalent Dosea
(mg/kg-day)
0
0.062
0.62
5.8
0
0.054
0.55
5.3
Renal Tubular Neoplasm
Incidence
1/90(1.1%)
0/40 (0%)
0/40 (0%)
9/39 (23%)
0/90 (0%)
0/40 (0%)
0/40 (0%)
6/40(15%)
a Human Equivalent Dose = Animal dose (Animal body weight/Human body weight)174
source: Kociba et al. (1977)
Increased renal tubular hyperplasia and renal tubular adenomas and adenocarcinomas (some
of which metastasized to the lungs), were found in rats exposed to 20 mg/kg-day of HCBD for up
to 2 years. Lesser degrees of toxicity, including an increase in renal tubular hyperplasia, were found
in rats ingesting 2 mg/kg-day for up to 2 years. A composite dose-related change in the rodent
kidney leading to tumor formation is shown in Table 8-4. This pattern is consistent with the
hypothesis that renal tumor formation may require, and be secondary to, renal cytotoxicity induced
by exposure to HCBD. The mode of action established for HCBD implies that, following exposures
less than those which produce overt renal damage, no significant excess carcinogenic risk can be
attributed to HCBD.
8.2.2 Dose-Response Characterization
The Kociba et al. (1977) and NTP (1991) studies were used to quantify the cancer risk from
ingested HCBD, as discussed below.
1986 Guidance: Linearized Multistage Model
The current IRIS file contains a carcinogenicity assessment of HCBD based on EPA's 1986
Guidelines for Carcinogen Risk Assessment (U. S. EPA, 1986). The dose-response data for male rats
HCBD February 2003
8-6
-------
were fitted to the linearized multistage model. To estimate human equivalent dose from an animal
study, the doses administered to animals were adjusted by a scaling factor of (body weight)273. The
resulting cancer slope factor is 7.8 x 10"2 (mg/kg-day)"1. This slope factor corresponds to a drinking
water unit risk of 2.2 x 10"6 per |ig/L (U.S. EPA, 1997a), and the drinking water concentration that
corresponds to a lifetime excess cancer risk of 1 x 10"6 is 0.5 |ig/L.
Table 8-4. Dose-Related Changes in the Rodent Kidney after Oral Exposure to HCBD,
Chronic Study - Rat (Kociba et al., 1977).
Dose (mg/kg-day)
terminal kidney
weight increase
(abs. & rel.)
hyperplasia - multi
focal
hyperplasia-
adenomatous
tumors
0.2
-
-
-
-
2
-
?
+
(? only)
-
20
+
+
+
+
1996,1999 Proposed Guidance
The draft Ambient Water Quality Criteria for hexachlorobutadiene (U.S. EPA, 1998a)
utilized the methodology discussed in EPA's 1996 Proposed Guidelines for Carcinogen Risk
Assessment (U.S. EPA, 1996a) to evaluate the carcinogenicity of HCBD. This has been updated in
the 1999 draft Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999). Under these
guidelines, two approaches can be used for dose-response extrapolation for quantification of cancer
risk, depending on what is known about the mode of action for carcinogenicity and the shape of the
dose-response curve. A linear approach is used for a chemical when available evidence indicates the
chemical has direct DNA mutagenic activity or is DNA-reactive, or when the evidence supports
another mode of action that is anticipated to be linear. An inference of linearity may also be
supported if existing human exposure is high and near doses associated with key events in the
carcinogenesis process. The linear approach is used as a matter of policy if there is an absence of
sufficient mode-of-action information on tumorigenesis. The nonlinear approach may be used when
the tumor mode-of-action supports nonlinearity (e.g., some cytotoxic and hormonal agents) and the
chemical does not demonstrate mutagenic effects consistent with linearity. The nonlinear approach
is also selected when a mode of action supporting nonlinearity has been demonstrated, and the
chemical has some indication of mutagenic activity, but is judged not to play a significant role in
tumor causation. As a matter of science policy, nonlinear probability functions are not fitted to tumor
response data to extrapolate quantitative low-dose risk estimates because different models can lead
HCBD February 2003
8-7
-------
to a wide range of results, and there is currently no basis to choose among them. In these cases, the
RfD is generally used for protection of cancer effects.
Because both linear (mutagenic) and nonlinear (toxicity associated) mode of action for
carcinogenicity of HCBD may be operating in vivo, both of these approaches were evaluated in the
draft Ambient Water Quality Criteria for hexachlorobutadiene (U. S. EPA, 1998a) for characterizing
the carcinogenic hazard of HCBD, as discussed below. DNA adduct formation in mouse kidney was
weak, even at an oral dose of 30 mg/kg. However, hypertrophy and regeneration of renal tubule
epithelial cells precede neoplastic effects in both dose (0.2 vs. 20 mg/kg-day) and time (subchronic
vs. chronic responses). This suggests that nonlinear effects may underlie or contribute to the
carcinogenic potential of HCBD, although a role for genotoxic mechanisms cannot confidently be
eliminated. Based on these considerations, the linear method for quantitative cancer risk assessments
of HCBD would be the default approach.
Linear Approach
Because there are limited data which suggest that HCBD might be genotoxic and mutagenic
(see Section 7.3.1), this approach is considered in the dose-response extrapolation for HCBD.
Under the proposed guidelines, the cancer risk from a chemical is assessed in two steps. The
first step involves curve-fitting of the cancer dose-response data within the observable range to
derive a point-of-departure (Pdp) (U.S. EPA, 1999). The dose at the point-of-departure is expressed
as the human equivalent dose. The dose that causes a 10% increase in extra risk is referred to as the
ED10. The point-of-departure is defined as the 95% lower confidence limit on the ED10, and is
referred to as the LED10. The second step in the process is linear extrapolation of the dose-response
curve from the LED10 to the origin, and determination of the slope of that line.
The LED10 for HCBD was calculated by fitting the quantal polynomial model7 to the tumor
dose response data reported by Kociba et al. (1977). Since the mortality rate was significantly
increased in the male rats exposed at the high dose (which is the only dose with an increased tumor
incidence in animals), the tumor data from the female rats were used. In accordance with current
guidance (U.S. EPA, 1992d, 1999), the human equivalent dose was calculated by assuming dose
equivalency based on body weight raised to the 3/4 power. The best fit to the data is shown in Figure
8-2. The ED10 was found to be 4.9 mg/kg-day, and the LED10 was 2.5 mg/kg-day. Linear
extrapolation from the LED10 to the origin yields a slope of 4 x 10"2 (mg/kg-day)"1.
This modeling was carried out using the Global 86 multistage model software.
HCBD February 2003
-------
Figure 8-2. Renal Tumor Dose Response Curves.
LJJ
0.30
0.25
0.20
0.15
0.10
0.05
0.00
Lower confidence
limit on ED10
(LED10)
ED10
2345
Human Equi\alent Dose (mg/kg-day)
Non-linear Approach
The non-linear approach is used when available data indicate that the dose-response curve
for tumor induction is not linear, and that cancer may not be the result of a direct DNA-damage
mechanism. As discussed previously, data from the study by Kociba et al. (1977) indicate that the
dose response curve is strongly non-linear, and that renal tumors only occur at HCBD doses that
cause frank renal toxicity and increased mortality. Therefore, tumor data from this study are not
considered suitable for dose-response extrapolation, and the non-linear approach should be
evaluated.
For HCBD, mode-of-action considerations suggest carcinogenicity may be secondary to
renal toxicity, which has a threshold (see Section 7.4.3, Mode of Action), and the RfD approach was
used for non-linear analysis in accordance with the proposed cancer guidelines (U.S. EPA, 1999).
The point-of-departure (Pdp) selected for use was the BMDL for renal tubular histological lesions
in female mice. This is because a Pdp based on a sensitive key precursor of the neoplastic response
is more protective and more reliable than a Pdp based on the neoplastic response itself (U.S. EPA,
1999). The BMDL for renal toxicity calculated from the NTP (1991) study was 0.1 mg/kg-day in
HCBD February 2003
-------
female mice. Application of a composite uncertainty factor of 300 (10 for extrapolation from
animals to humans; 10 for human variability; and 3 for database deficiencies) yields a dose of
3 x 10"4 mg/kg-day, which is the RfD for HCBD. This RfD is also protective of potential
carcinogenic effects under the non-linear approach.
By the non-linear approach, a health reference level (HRL) of 2 |ig/L can also be derived using the
Reference Dose (RfD) for HCBD of 3 x 10'4 mg/kg-day. The RfD is an estimate of the daily oral
dose to the human population that is likely to be without appreciable risk of adverse effects over a
lifetime exposure. This dose was converted to a drinking water equivalent concentration of 10 |ig/L
by multiplying the RfD by the default body weight for an adult (70 kg) and dividing the result by
the default daily intake of drinking water for an adult (2 L/day). For derivation of the HRL, it was
assumed that about 20% of an individual's total exposure to HCBD was attributable to drinking
water. Multiplication of the drinking water equivalent concentration by 0.2 yields the HRL of 2
l-ig/L (rounded to 1 significant number). The HRL was derived as follows:
HRL = RfD x BW x RSC
DI
Where:
RfD = Reference dose for HCBD in drinking water, 3 x 10"4 mg/kg-
day
BW = Body weight of an adult, 70 kg
DI = Daily intake of water for an adult, 2 L/day
RSC = Relative Source Contribution, default value of 20%
Therefore:
HRL = (3 x 1Q-4 mg/kg-dav^) x ( 70 ke.} x p.20
2 L/day
= 2 |ig/L (rounded to 1 significant number).
However, the HRL of 0.9 |ig/L derived from linear approach is used as the preliminary health
effect level in this document. (See Section 8.2.5). The linear approach is used as the default because
of potential genotoxicity of HCBD metabolites.
8.2.3 Extrapolation Model and Rationale
In Section 8.2.2., the carcinogenicity of HCBD was evaluated using both linear and non-
linear approaches. Because of the lack of data, it is not certain which method of cancer risk
evaluation is most appropriate for HCBD. On the one hand, some tests indicate that one or more of
the metabolites of HCBD are mutagenic, suggesting direct damage to renal mitochondrial DNA by
its reactive metabolites. On the other hand, direct observations on cancer dose-response clearly
support a nonlinear curve, with no observable increase in tumors at doses that do not induce
significant renal necrosis and regeneration. This is supported by the observation that tumors occur
HCBD February 2003 8-10
-------
only in the kidney and not in other tissues that are not significantly injured by HCBD. Therefore,
the tumor data from the Kociba et al. (1977) study are not considered suitable for linear dose-
response extrapolation.
Although HCBD metabolites have some indication of mutagenic activity, they are not likely
to play a significant role in tumor causation due to their weak activity. Moreover, mutations of
mitochondrial DNA may result in mitochondrial dysfunction (See Section 7.4.3), which would
support cytotoxicity and nonlinear approach. There should also be decreased concern over
genotoxicity for humans because the activity of HCBD metabolizing enzymes, particularly renal P-
lyase, may be many fold lower in humans than the corresponding enzymes in rats (see section 6.3).
In addition, human exposure levels (see Section 9.3) are about 4 orders of magnitude lower than the
human equivalent dose corresponding to the dose at which tumor incidence was reported in Kociba
et al. (1977) study. In consideration of the overall evidence, the non-linear approach may be
appropriate for HCBD. The draft Ambient Water Quality Criteria Document for
Hexachlorobutadiene (U.S. EPA, 1998a) has recommended using the nonlinear approach for
carcinogenicity assessment of HCBD.
According to the 1999 U.S. EPA Draft Guidelines for Carcinogen Risk Assessment, "When
the mode of action information indicates that the dose-response may be adequately described by
both a linear and a nonlinear approach, then the default is to present both the linear and margin of
exposure analyses." As this can be done without contradicting previous guidelines, both linear and
nonlinear analyses are presented.
8.2.4 Cancer Potency and Unit Risk
Table 8-5 summarizes the cancer values derived for HCBD. Analysis of tumor dose-
response information from the Kociba et al. (1977) study using the linear extrapolation approach
from the proposed carcinogen risk assessment guidelines (U.S. EPA, 1999) resulted in a slope factor
of 4 x 10"2 (mg/kg-day)"1. This value is about half of the slope factor of 7.8 x 10"2 (mg/kg-day)"1
derived previously using the linearized multistage (LMS) model (U. S. EPA, 1997a), but most of the
apparent difference may be attributable to the different methods used to calculate human equivalent
doses from the animal doses (the scaling factor used in the LMS approach assumed body weight to
the 2/3 power, while a factor of body weight raised to the 3/4 was used for the Pdp method). Based
on the slope factor of 4 x 10"2 (mg/kg-day)"1 derived using the LED10 approach with linear
extrapolation, the unit risk is 1.1 x 10"6 per (i-ig/L) and the drinking water concentration that
corresponds to a lifetime excess risk of 1 x 10"6 is 0.9 |ig/L.
HCBD February 2003 8-11
-------
Table 8-5. Summary of Cancer Risk Values for HCBD.
Approach
LMSa
(U.S. EPA,
1991)
LED10, linear
extrapolation
Nonlinear
Parameter
Slope
Unit Risk
Water Concentration at
risk of 1 x 1Q-6
LED10 (tumors)b
Slope
Unit Risk
Water Concentration at
risk of 1 x 1Q-6(HRL)
Pdp (BMDL)
Uncertainty factor
HRL
Value
7.8 x ID'2 (mg/kg-day)-1
2.2 x 1Q-6 per(ng/L)
0.5 jig/L
2.5 mg/kg-day
4x 1Q-2 (mg/kg-day)-1
1.1 x 1Q-6 per(|ig/L)
0.9 jig/L
0.1 mg/kg-day
300
2[igfL
a Animal to human dose extrapolation based on body weight273
b Animal to human dose extrapolation based on body weight374
8.2.5 Discussion of Confidence
The available database associating HCBD and carcinogen!city is limited. There are no human
data. The evidence is obtained only in one chronic dietary study in a single species (Sprague-Dawley
rats) (Kociba et al., 1977), where rats developed severe renal toxicity preceding tumor formation.
The tumors were seen only at a high dose which exceeded the maximum tolerated dose (MTD, i.e.,
greater than 10% body weight depression) in both sexes of rats and produced high mortality in the
males. Similar renal toxicity observed in a 30-day study of HCBD in rats by the same laboratory and
in another 90-day subchronic study in mice (NTP, 1991) strengthens the idea that the tumor
formation is induced by cytotoxicity. Both the NTP (1991) and Kociba et al. (1977) studies tested
a sufficient number of animals.
A limitation of the Kociba et al. (1977) study is the selection and spacing of doses. Although
the study employed an adequate number of animals, the doses selected for testing were separated
by a factor of 10 (0, 0.2, 2, and 20 mg/kg-day). Thus, there are no observations between the dose of
2 mg/kg-day (causing no tumors), and the dose of 20 mg/kg-day (causing a 15% tumor response
in females and a 23% tumor response in males). More doses between 2 and 20 mg/kg-day would
better delineate the shape of the dose-response curve.
A weight-of-evidence analysis of the available data as a whole indicates that the confidence
in using either the linear or nonlinear approach is not high; this is particularly true for the linear
HCBD February 2003
-------
method which is based on only one data point at a high-dose exceeding the MTD. However, the
possible genotoxicity of HCBD metabolites forces the use of the linear approach as default. Thus,
the HRL of 0.9 |ig/L derived from the linear approach is used as the preliminary health effect level
in this document.
HCBD February 2003 8-13
-------
9.0 REGULATORY DETERMINATION AND CHARACTERIZATION OF RISK FROM
DRINKING WATER
9.1 Regulatory Determination for Chemicals on the CCL
The Safe Drinking Water Act (SDWA), as amended in 1996, required the Environmental
Protection Agency (EPA) to establish a list of contaminants to aid the Agency in regulatory priority
setting for the drinking water program. EPA published a draft of the first Contaminant Candidate
List (CCL) on October 6, 1997 (62 FR 52193, U.S. EPA, 1997b). After review of and response to
comments, the final CCL was published on March 2, 1998 (63FR 10273, U.S. EPA 1998d). The
CCL grouped contaminants into three major categories as follows:
Regulatory Determination Priorities - Chemicals or microbes with adequate data to support
a regulatory determination,
Research Priorities - Chemicals or microbes requiring research for health effects, analytical
methods, and/or treatment technologies,
Occurrence Priorities - Chemicals or microbes requiring additional data on occurrence in
drinking water.
The March 2, 1998 CCL included one microbe and 19 chemicals in the regulatory
determination priority category. More detailed assessments of the completeness of the health,
treatment, occurrence and analytical method data led to a subsequent reduction of the regulatory
determination priority chemicals to a list of 12 (one microbe and 11 chemicals) which was
distributed to stakeholders in November 1999.
SDWA requires EPA to make regulatory determinations for no fewer than five contaminants
in the regulatory determination priority category by August 2001. In cases where the Agency
determines that a regulation is necessary, the regulation should be proposed by August 2003 and
promulgated by February 2005. The Agency is given the freedom to also determine that there is no
need for a regulation if a chemical on the CCL fails to meet one of three statutory criteria established
by SDWA and described in Section 9.1.1.
9.1.1 Criteria for Regulatory Determination
These are the three criteria used to determine whether or not to regulate a chemical on the
CCL:
The contaminant may have an adverse effect on the health of persons,
The contaminant is known to occur, or there is a substantial likelihood that the contaminant
will occur, in public water systems with a frequency and at levels of public health concern,
In the sole judgment of the administrator, regulation of such contaminant presents a
meaningful opportunity for health risk reduction for persons served by public water systems.
HCBD February 2003 9-1
-------
The findings for all criteria are used in making a determination to regulate a contaminant.
As required by SDWA, a decision to regulate commits the EPA to publication of a Maximum
Contaminant Level Goal (MCLG) and promulgation of a National Primary Drinking Water
Regulation (NPDWR) for that contaminant. The agency may determine that there is no need for a
regulation when a contaminant fails to meet one of the criteria. A decision not to regulate a
contaminant is considered a final Agency action and is subject to judicial review. The Agency can
choose to publish a Health Advisory (a nonregulatory action) or other guidance for any contaminant
on the CCL independent of the regulatory determination.
9.1.2 National Drinking Water Advisory Council Recommendations
In March 2000, the EPA convened a Working Group under the National Drinking Water
Advisory Council (NDWAC) to help develop an approach for making regulatory determinations.
The Working Group developed a protocol for analyzing and presenting the available scientific data,
and recommended methods to identify and document the rationale supporting a regulatory
determination decision. The NDWAC Working Group report was presented to and accepted by the
entire NDWAC in July 2000.
Because of the intrinsic difference between microbial and chemical contaminants, the
Working Group developed separate but similar protocols for microorganisms and chemicals. The
approach for chemicals was based on an assessment of the impact of acute, chronic, and lifetime
exposures, as well as a risk assessment that includes evaluation of occurrence, fate, and dose-
response. The NDWAC protocol for chemicals is a semi-quantitative tool for addressing each of the
three CCL criteria. The NDWAC requested that the Agency use good judgement in balancing the
many factors that need to be considered in making a regulatory determination.
The EPA modified the semi-quantitative NDWAC suggestions for evaluating chemicals
against the regulatory determination criteria and applied them in decision making. The quantitative
and qualitative factors for hexachlorobutadiene (HCBD) that were considered for each of the three
criteria are presented in the sections that follow.
9.2 Health Effects
The first criterion asks if the contaminant may have an adverse effect on the health of
persons. Because all chemicals have adverse effects at some level of exposure, the challenge is to
define the dose at which adverse health effects are likely to occur, and estimate a dose at which
adverse health effects are either not likely to occur (threshold toxicant), or have a low probability
for occurrence (non-threshold toxicant). The key elements that must be considered in evaluating the
first criterion are the mode of action, the critical effect(s), the dose-response for critical effect(s), the
RfD for threshold effects, and the slope factor for non-threshold effects.
A description of the health effects associated with exposure to HCBD is presented in Chapter
7 of this document and summarized below in Section 9.2.2. Chapter 8 and Section 9.2.3 present
dose-response information, where applicable, for threshold and non-threshold health effects.
HCBD February 2003 9-2
-------
9.2.1 Health Criterion Conclusion
The available toxicological data indicate that HCBD has the potential to cause adverse health
effects in animals, and probably in humans. The available human data involve inhalation exposure
and are confounded by simultaneous exposures to other chemicals in an occupational setting; thus,
attributing observed effects to specific levels of HCBD exposure is not possible. In rodents, there
is clear evidence of renal damage resulting from acute, subchronic, and chronic HCBD oral
exposures. A few animal studies have also reported liver effects and neurotoxicity. Review of
animal dose-response data endpoints indicates that subchronic and chronic LOAEL values for
HCBD toxicity are generally at 2 mg/kg-day and above. The RfD for HCBD is 3 x 10"4 mg/kg-day
(Chapter 8). Limited evidence of carcinogenic potential in rodents suggests that HCBD may be
carcinogenic secondary to renal tubular epithelial cell cytotoxicity. However, data in humans are
lacking. Both linear and non-linear approaches were evaluated for cancer dose-response assessment.
Using the non-linear approach, the RfD protects both non-cancer and cancer effects. However, in
the presence of data supporting the potential genotoxicity of HCBD metabolites, the linear approach
is used as default, with a 10"6 risk at a drinking water concentration of 0.9 |ig/L.
9.2.2 Hazard Characterization and Mode of Action Implications
Data for the human health effects of HCBD are limited to a few studies of occupational
exposure to HCBD. A relationship could not be established from these studies between HCBD
exposure and toxic effects either because of concurrent exposure to other chemicals or because of
equivocal results.
Studies in animals show the selective effect of HCBD on the kidney, specifically the
proximal tubule. Renal toxicity in rodents has been shown with single acute exposures to 100-200
mg HCBD/kg, and with short-term exposures to 3 mg/kg-day and above. Subchronic and chronic
studies in rodents show clear dose-related renal damage at 2 mg/kg-day and above. Progressive
events overtime include changes in kidney weight, increased urinary excretion of coproporphyrin,
and increased renal tubular epithelial hyperplasia.
Other noncancer effects associated with HCBD exposure in animals include developmental
effects and neurotoxicity (Harleman and Seinen, 1979; Badaeva, 1983; Badaeva et al., 1985).
However, these effects were observed at higher doses than for renal toxicity. Pups with lower birth
weights and reduced growth were reported at maternal dose of 8.1-15 mg/kg-day in rats (Badaeva,
1983; Harleman and Seinen, 1979).
Results from mutagenicity studies with HCBD are ambiguous. In the presence of appropriate
metabolic activation conditions, HCBD and its metabolites are mutagenic in some (Vamvakas et al.,
1988; Reichert et al., 1984), but not all studies. HCBD metabolites have been shown to bind to
mitochondrial DNA in vivo in mice (Schrenk and Dekant, 1989), and induce DNA repair in cultured
porcine kidney cells (Vamvakas et al., 1989), suggesting its genotoxic potential. No human studies
of HCBD carcinogenicity have been reported and only one lifetime animal study has been performed
(Kociba et al., 1977). In this study, neoplastic changes occurred only at the highest dose which
exceeded the maximum tolerated dose (MTD), i.e. there was increased mortality, greater than 10%
HCBD February 2003 9-3
-------
decrease in body weight and severe renal toxicity. Because these significant adverse effects were
observed at the high dose, tumor formation may be secondary to cytotoxicity.
The nephrotoxicity of HCBD is dependent on a multistep bioactivation mechanism
involving both kidney and liver enzymes. The ultimate step in this pathway is a P-lyase mediated
degradation of a HCBD metabolite that generates a highly reactive thioketene in proximal tubule
cells.. In vitro studies suggest that cortical mitochondria are the critical subcellular target for toxicity
of the bioactivated sulfur conjugates of HCBD. Covalent binding of this reactive HCBD metabolite
to cellular macromolecules (e.g. proteins, mitochondrial DNA), and the resultant mitochondrial
dysfunction is believed to contribute to the renal cytotoxicity and tumors observed in animals.
Recent evidence suggests that HCBD-induced a2(J-globulin accumulation contributes to renal injury
in male rats. However, renal tubular necrosis and renal tubular tumors were observed in both male
and female rats following HCBD exposure (Kociba et al., 1977), and renal necrosis and regeneration
were also observed in male and female mice (NTP, 1991). Therefore, a2(J-globulin accumulation
cannot be the sole mechanism for HCBD-induced carcinogenesis.
One important issue in the evaluation of the hazard posed by HCBD is the applicability of
rodent mechanistic data to humans. In vitro studies with human renal cytosol and cultured human
proximal tubule cells suggest that humans have the potential to form the HCBD-glutathione
conjugates and to metabolize HCBD cysteine conjugates to toxic metabolites. However, the rate of
metabolism, particularly for the reaction catalyzed by P-lyase, appears to be much lower for humans
than rodents (Lock, 1994; Lash et al., 1990).
It has been generally observed that existing nephropathy or age-related kidney degeneration
can increase the risk of renal injury or exacerbate nephrotoxicity in humans. Therefore, sensitive
populations for HCBD exposure may include people with pre-existing kidney or liver damage or the
elderly. Although it is unlikely that human newborns would be acutely exposed to significant doses
of HCBD, acute exposures for young rats and mice cause toxicity at lower doses than for adults
(Hook et al., 1983; Lock et al., 1984).
9.2.3 Dose-Response Characterization and Implications in Risk Assessment
Dose-response information from several key studies of HCBD toxicity in animals is
summarized in Table 9-1. These studies currently provide the most reliable information on threshold
levels for HCBD toxicity in animals exposed via the oral route.
Noncancer effects
In short-term studies, a LOAEL of 10 mg/kg-day and a NOAEL of 3 mg/kg-day were
identified for reduced body weight gain and food consumption in female Sprague-Dawley rats
administered HCBD in their diets for 30 day. Renal tubular degeneration, necrosis and regeneration
HCBD February 2003 9-4
-------
Table 9-1. Dose-Response Information from Several Key Studies of HCBD Toxicity (Oral
Exposure).
Study
Species
No./
Sex
Doses
mg/kg-day
Duration
NOAEL
mg/kg-
day
LOAEL
mg/kg-
day
Effects
Short-term Studies
Kociba et al.
(1977)
Jonker et al.
(1993b)
Harleman
and Seinen
(1979)
Stott et al.
(1981)
NTP (1991)
Rat
Sprague-
Dawley
Rat
Wistar
Rat
Wistar
Rat
Sprague-
Dawley
Mouse
B6C3F!
4F
5M
5F
6M
6F
5M
5M
5F
1.310e+10
2.25828
0
4.6
14.0
35.3
0.22
OM OF
3M 5F
12 M 16 F
40 M 49 F
30 days
4 weeks
14 days
3 weeks
2 weeks
3
2.25
0.2
..
10
8
4.6
20
3M 5F
Reduced body
weight gain, food
consumption;
increased
hemoglobin
concentration,
relative kidney
weight; renal
tubular
degeneration,
necrosis,
regeneration.
Decreased liver
weight, plasma
creatinine, body
weight, adrenal
weight; renal
tubular
cytomegaly.
Decreased body
weight gain and
food conversion
efficiency; renal
tubular epithelial
cell degeneration.
Decreased body
weight gain;
increased relative
kidney weight;
kidney damage.
Renal tubular
necrosis.
HCBD February 2003
9-5
-------
Table 9-1 (continued)
Study
Species
No./
Sex
Doses
mg/kg-day
Duration
NOAEL
mg/kg-
day
LOAEL
mg/kg-
day
Effects
Subchronic Studies
NTP (1991)
Mouse
B6C3FJ
10
M
10 F
OM OF
0.1 M 0.2 F
0.4 M 0.5 F
1.5M 1.8F
4.9 M 4.5 F
16.8 M 19.2 F
13 weeks
1.5 M
0.2 F
4.9 M
0.5 F
Renal tubular cell
regeneration
(increased
epithelial nuclei
and basophilic
staining)
Chronic Studies
Kociba et al.
(1977)
Rat
Sprague-
Dawley
39-
40
M, F
0.222
22-24
months
0.2
2
Increased kidney
weight; renal
tubular epithelial
hyperplasia and
neoplasia.
M = male; F = female
were observed at 30 mg/kg-day (Kociba et al., 1971; Schwetz et al., 1977). A LOAEL of 8 mg/kg-
day and a NOAEL of 2.25 mg/kg-day were identified for decreased body weight gain and renal
tubular effects in Wistar rats given HCBD in their diets for 4 weeks (Jonker et al., 1993b). A 3-week
oral exposure with male Sprague-Dawley rats identified a LOAEL of 20 mg/kg-day and a NOAEL
of 0.2 mg/kg-day for kidney damage and increased relative kidney weight (Stott et al., 1981), and
a 2-week feeding study in Wistar rats identified a LOAEL of 4.6 mg/kg-day (the lowest dose tested)
for renal tubular epithelial cell degeneration (Harleman and Seinen, 1979). A 2-week oral exposure
study in B6C3FJ mice reported a LOAEL of 3-5 mg/kg-day (the lowest dose tested) for renal tubular
necrosis (NTP, 1991). Thus, renal effects in rodents resulting from short-term exposure to HCBD
appear to have LOAELs of around 5-20 mg/kg-day, depending on the species and strain used, the
length of exposure, and the method of administration.
In a subchronic oral exposure study of HCBD inB6C3F1 mice, a NOAEL of 1.5 mg/kg-day
was identified for male mice based on renal tubular cell regeneration (NTP, 1991). Tubular
regeneration occurred in 1 of 10 females in the lowest dose group (0.2 mg/kg-day). The study
authors concluded that a NOAEL for female mice could not be identified from these data (NTP,
1991). However, EPA (U.S. EPA, 1998a) and others (WHO, 1994) have concluded that the effect
observed at 0.2 mg/kg-day is not statistically significant, and therefore consider this dose to be the
NOAEL for female mice. Because tubular regeneration occurred in 1 of 10 females at 0.2 mg/kg-
day, this NOAEL may be close to a minimal LOAEL for renal injury.
Only one study of lifetime oral exposure to HCBD was located (Kociba et al., 1977). This
study identified a NOAEL of 0.2 mg/kg-day and a LOAEL of 2 mg/kg-day in rats, based on an
HCBD February 2003
9-6
-------
increase in renal tubular epithelial cell hyperplasia/regeneration. The value of this NOAEL from a
chronic study is the same as the equivocal NOAEL of 0.2 mg/kg-day identified in the 13-week NTP
study in female mice (NTP, 1991), indicating the female mice may be more sensitive than rats to
HCBD.
The Reference Dose (RfD) for HCBD is 3 x 10'4 mg/kg-day (Chapter 8). The RfD is "an
estimate (with uncertainty spanning approximately an order of magnitude) of a daily exposure to the
human population (including sensitive subgroups) that is likely to be without appreciable risk of
deleterious effects over a lifetime" (U.S. EPA, 1993). The RfD is derived from a BMDL of 0.1
mg/kg-day for renal tubular epithelial cell hyperplasia/regeneration from the NTP (1991) studies.
A composite uncertainty factor of 300 was used in the derivation of the RfD to account for:
extrapolation from animals to humans (factor of 10); protection of sensitive subpopulations (factor
of 10); and database deficiency (factor of 3) because of lack of a 2-generation reproductive study.
Cancer effects
The single lifetime exposure study in rats is also a source of data on tumor formation (Kociba
et al., 1977). Only at the highest dose, 20 mg/kg-day, were tumors seen in both sexes. This dose
exceeded the level at which significant noncancer effects were seen, such as mortality, renal toxicity,
and body weight depression. In this study, the second highest dose was 2 mg/kg-day and there were
no tumors in this exposed group. While slope of the dose-response curve cannot be determined from
the data set, it must be kept in mind that the purpose of the Kociba et al.(1977) bioassay was for
hazard identification, not quantitative risk analysis.
Under EPA's 1986 Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1986), HCBD
is classified as Group C, possible human carcinogen. Using the linearized multistage model, a slope
factor of 7.8 x 10"2 per mg/kg-day was calculated at the 95th upper confidence level (U.S. EPA,
1991c). Under EPA's 1999 draft Guidelines for Carcinogen Risk Assessment (USEPA, 1999),
HCBD is classified as likely to be carcinogenic to humans by the oral route of exposure, but whose
carcinogenic potential by the inhalation and dermal routes of exposure cannot be determined
because there are inadequate data to perform an assessment. Both the linear and nonlinear dose-
response extrapolation approaches were used to quantify cancer risk (U.S. EPA, 1998a) because
both cytotoxicity and mutagenic mode of action may be involved. The linear approach yields a slope
of 4 x 10"2 per mg/kg-day. Using the non-linear approach, the RfD was used, yielding the resulting
dose of 3 x 10"4 mg/kg-day. EPA's draft Ambient Water Quality Criteria for hexachlorobutadiene
(U.S. EPA, 1998a) recommended using the non-linear approach for dose-response extrapolation. As
discussed previously, data from Kociba et al. (1977) indicated that the tumor dose response curve
is strongly non-linear, and that renal tumors only occur at HCBD doses that cause frank toxicity.
However, the presence of data supporting the genotoxicity of HCBD metabolites forces the
conservative choice of a linear approach to be more appropriate for HCBD.
The conclusion from the dose response analysis is that HCBD is a weak carcinogen because
it is carcinogenic only at cytotoxic dose.
HCBD February 2003 9-7
-------
9.3 Occurrence in Public Water Systems
The second criterion asks if the contaminant is known to occur or if there is a substantial
likelihood that the contaminant will occur in public water systems with a frequency and at levels of
public health concern. In order to address this question, the following information was considered:
Monitoring data from public water systems
Ambient water concentrations and releases to the environment
Environmental fate
Data on the occurrence of HCBD in public drinking water systems were the most important
determinants in evaluating the second criterion. EPA looked at the total number of systems that
reported detections of HCBD, as well as those that reported concentrations of HCBD above an
estimated drinking water health reference level (HRL) (U.S. EPA, 200Ic). For noncarcinogens, the
estimated HRL level was calculated from the RfD assuming that 20% of the total exposure would
come from drinking water. For carcinogens, the HRL was the 10"6 risk level. The HRLs are
benchmark values that were used in evaluating the occurrence data while the risk assessments for
the contaminants were being developed.
The available monitoring data, including indications of whether or not the contamination is
a national or a regional problem, are included in Chapter 4 of this document and summarized below.
Additional information on production, use, and fate are found in Chapters 2 and 3.
9.3.1 Occurrence Criterion Conclusion
HCBD has never been specifically manufactured as a commercial product in the United
States, but is generated as waste by-product from the chlorination of hydrocarbons. The available
data for HCBD use indicate an overall downward trend. The ten-year pattern of TRI releases to
surface water is variable but generally decreasing within the range from 5 to 1,911 pounds. The
physicochemical properties of HCBD and the available data for environmental fate indicate that
HCBD in surface water is likely to be rapidly degraded by biotic and abiotic processes although it
has the potential for bioaccumulation. Monitoring data indicate that HCBD is infrequently detected
in public water supplies. When HCBD is detected, it very rarely exceeds the HRL or a value of one-
half of the HRL. Chemical treatment of drinking water and leaching from drinking water surfaces
are not expected to contribute to significantly elevated levels of HCBD in drinking water.
9.3.2 Monitoring Data
Drinking Water
HCBD has been detected in a small percentage of public water supply (PWS) samples
collected under the authority of the Safe Drinking Water Act. Occurrence data for HCBD in drinking
water are presented and analyzed in Chapter 4 of this document. Data from two monitoring periods
were available for analysis. Data from Round 1 were collected during the period 1987 to 1992. Data
HCBD February 2003 9-8
-------
from Round 2 were collected during the period 1993 to 1997. Round 1 and 2 monitoring detected
HCBD in only 0.13% and 0.05% of all samples analyzed, respectively. When data are expressed on
a PWS basis, Round 1 and Round 2 monitoring detected HCBD at least once in 0.35% (228 systems)
and 0.18% (117 systems) of the tested water supplies, respectively.
The median and 99th percentile concentrations for all samples (i.e., samples with and without
detectable levels of HCBD) were below the minimum reporting level (MRL). When subsets of the
data containing only samples with detectable levels of HCBD were analyzed, the median and 99th
percentile concentrations for Round 1 were 0.25 |ig/L and 10 |ig/L, respectively. The median and
99th percentile for Round 2 detections were 0.30 |ig/L and 1.5 |ig/L, respectively.
When monitoring results were compared to a value of one-half of the HRL, 0.163% of Round
1 (106 systems) and 0.079% of Round 2 (51 systems) water supplies exceeded this benchmark at
least once during the reporting period. The percentages of water supplies that exceeded the HRL at
least once in Round 1 andRound2monitoringwere0.114%(74systems)and0.018%(ll systems),
respectively.
PWSs with detected levels of HCBD were widely distributed throughout the United States
(see Figures 4-2 and 4-3 in this document), and no clear patterns of regional geographic occurrence
were evident.
Ambient Water
HCBD has not been detected in the ground water samples reviewed and/or analyzed under
the U. S. Geological Survey National Ambient Water Quality Assessment (NAWQA) program. The
first round of intensive monitoring in the ongoing NAWQA was conducted from 1991 to 1996 and
targeted 20 watersheds. Data from each NAWQA study unit were augmented by additional data
from local, state, and federal agencies that met specified criteria.(See Section 4.1.1). HCBD was not
detected in rural and urban wells of the local, State, and federal data set compiled by NAWQA.
These data represent untreated ground water of the conterminous United States for the years 1985-
1995.
A review of highway and urban runoff studies also found no detections of HCBD.
9.3.3 Use and Fate Data
Significant quantities of HCBD are generated in the United States as waste by-product from
the chlorination of hydrocarbons, although HCBD has never been specifically manufactured as a
commercial product domestically. No recent estimate could be found on the by-product amounts,
but in 1982, it was estimated that about 28 million pounds were generated (ATSDR, 1994). HCBD
imports dropped during the late 1970s, the period for which data are reported (Howard, 1989).
In all environmental media, HCBD binds strongly to particles (ATSDR, 1994). It is readily
adsorbed to airborne particulate matter, to sediments in water, and to soil organic particles.
Volatilization from soil or water to air appears to occur relatively slowly (U.S. EPA, 1991a).
HCBD February 2003 9-9
-------
Very little information is available on degradation or transformation of HCBD. Under
aerobic conditions, HCBD in sewage contaminated waters showed complete biodegradation (Tabak
et al., 1981). Under anaerobic soil conditions, biodegradation will not occur based on results
obtained in sludge incubated under anaerobic conditions (Johnson and Young, 1983). Estimates of
the half-life of HCBD in water range from 3 to 30 days in rivers and from 30 to 300 days in lakes
and ground water (Zoeteman et al., 1980).
HCBD may readily partition from the water into biological tissues, as suggested by its high
log octanol:water partition coefficient (Kow of 4.78). Laboratory and field studies have confirmed
its bioaccumulation potential (WHO, 1994; U.S. EPA, 1998a). There is no evidence that HCBD has
biomagnification potential (WHO, 1994).
HCBD is not used as a drinking water treatment chemical, and leaching from drinking water
contact surfaces is not likely. Therefore, these factors are not expected to contribute to elevated
levels of HCBD in drinking water.
9.4 Risk Reduction
The third criterion asks if, in the sole judgment of the Administrator, regulation presents a
meaningful opportunity for health risk reduction for persons served by public water systems. In
evaluating this criterion, EPA looked at the total exposed population, as well as the population
exposed above the estimated HRL. Estimates of the populations exposed and the levels to which
they are exposed were derived from the monitoring results. These estimates are included in Chapter
4 of this document and summarized in Section 9.4.2 below.
In order to evaluate risk from exposure through drinking water, EPA considered the net
environmental exposure in comparison to the exposure through drinking water. For example, if
exposure to a contaminant occurs primarily through ambient air, regulation of emissions to air
provides a more meaningful opportunity for EPA to reduce risk than does regulation of the
contaminant in drinking water. In making a regulatory determination, the available information on
exposure through drinking water (Chapter 4) and information on exposure through other media
(Chapter 5) were used to estimate the fraction that drinking water contributes to the total exposure.
The EPA findings are discussed in Section 9.4.3 below.
In making its regulatory determination, EPA also evaluated effects on potential sensitive
populations, including the fetus, infants and children. Sensitive population considerations are
included in section 9.4.4.
9.4.1 Risk Reduction Criterion Conclusion
Approximately 2 to 5 million people are served by systems with detections of HCBD. An
estimated 10,000 of these individuals may be served by systems with detections greater than the
HRL, based on Round 2 monitoring data. Sensitive populations to HCBD may include people with
preexisting kidney damage and infants, though it is unlikely for human newborns to be acutely
exposed to significant doses of HCBD. When average daily intakes from drinking water are
compared with intakes from air, drinking water accounts for a relatively small proportion of total
HCBD February 2003 9-10
-------
HCBD intake. Relative intake rates from food may be higher, however, and intakes from soil are not
known. On the basis of these observations, the impact of regulating HCBD concentrations in
drinking water on health risk reduction is likely to be small.
9.4.2 Exposed Population Estimates
National population estimates for HCBD exposure were derived using summary statistics
for Round 1 and Round 2 PWS cross-sectional data (see Table 4-2 of this document) and population
data from the Water Industry Baseline Handbook (U.S. EPA, 2000e). Summary data for exposed
population estimates are provided in Table 9-2 below. An estimated 1.9 to 5 million people are
served by PWSs that have detected HCBD. Of this population, approximately 1.2 million people
could be exposed at one-half of the HRL, based on data from Round 1 sampling; and about 5 million
people could be exposed to over one-half the HRL, based on Round 2 sampling. Based on the data
from Round 1 sampling, about 781,000 individuals were exposed to concentrations at or above the
HRL. Based on Round 2 sampling results, an estimated 10,000 persons could be exposed at or above
the HRL. The Round 2 based estimate is probably a better estimate of possible exposure since the
database is more recent, and more representative of the cross-section population served by
groundwater.
Table 9-2. National Population Estimates for HCBD Exposure via Drinking Water.
Population of Concern
Served by PWS with detections
Served by PWSs with detections > (1/2
HRL)
Served by PWSs with detections > HRL
Round 1
1,909,000
1,213,000
781,000
Round 2
5,027,000
4,965,000
10,000
Source: Data taken from Table 4-2 of this document.
HRL = Health Reference Level
9.4.3 Relative Source Contribution
Relative source contribution analysis compares the magnitude of exposure expected via
drinking water to the magnitude of exposure from intake of HCBD in other media, such as food, air,
and soil. To perform this analysis, intake of HCBD from drinking water must be estimated.
Occurrence data for HCBD in water and other media are presented in Chapter 4 and 5 of this
document.
As shown in Table 4-2, the 99th percentile concentration for all samples (i.e., those with
detectable and nondetectable levels of HCBD) from Round 1 and Round 2 PWS sampling is below
the MRL. As a convention, a value of half the MRL is often used as an estimate of the concentration
of a contaminant in samples/systems whose results are less than the MRL. However, for Round 1
HCBD February 2003 9-11
-------
and Round 2, States have reported a wide range of values for the MRLs (See Section 4.2.1), and a
single estimate of the MRL for HCBD is unavailable.
As an alternative, the median concentration (0.3 |ig/L) for HCBD in samples with detectable
levels from both rounds was used to estimate intake from drinking water. The exposure estimate for
an average individual is determined by multiplying the drinking water concentration by daily water
intake (2 liters/day) and dividing by average adult body weight (70 kg), and is estimated to be 8.6
x 10"6 mg/kg-day. For children, assuming a daily water intake of 1 liter/day and body weight of 10
kg, the exposure estimate is 3.0 x 10"5mg/kg-day.
The estimated average daily intakes of HCBD from drinking water (based on median
concentration of detected samples) and other sources are shown in Table 9-3. The estimated
food:drinking water exposure ratio is 0.03 for an adult and 0.02 for a child (Table 9-4). The
estimated airdrinking water exposure is 14 for an adult and 21 for a child. Collectively, these data
indicate that intake from drinking water is low when compared to intake from air, though not
necessarily when compared to possible intake from food.
Table 9-3. Comparison of Average Daily Intakes from Drinking Water and Other Media a.
Medium
Drinking Water b
Food
Air
Adult (ng/kg-day)
8.6
0.15
120
Child (ng/kg-day)
30
0.24
630
a See Chapter 5 for derivation of intakes from media other than water
b Based on half the median values for detected hexachlorobutadiene concentrations
in Round 1 and Round 2
Table 9-4. Ratiosa of Exposures from Various Media to Exposures from Drinking Water.
Exposure Ratio
Food:Drinking Water
AirDrinking Water
Adult
0.02
14
Child
0.008
21
a Calculated from estimated daily intakes in Table 9-3
9.4.4 Sensitive Populations
The target organ for HCBD is primarily the kidney. Sensitive populations to HCBD exposure
may include people with preexisting kidney damage. Though it is unlikely that human newborns
would be acutely exposed to significant doses of HCBD, acute exposure for young rats causes
toxicity at lower levels than for adults (Hook et al., 1983; Lock et al., 1984).
HCBD February 2003
9-12
-------
Calculation of medium-specific exposure ratios (Table 9-4) indicates thatHCBD intake from
air is about 14- 20 fold greater than intake from water. Therefore, regulation of HCBD in drinking
water would be unlikely to significantly reduce the risk to sensitive populations.
9.5 Regulatory Determination Summary
While there is evidence that HCBD may have adverse health effects in humans at moderate-
to-high doses, it is unlikely that: 1) this contaminant will occur with a frequency or at concentrations
that are of public health concern; or 2) regulation of this contaminant represents a meaningful basis
for health risk reduction in persons served by public water systems. For these reasons, EPA does
not plan to regulate HCBD with a NPDWR..
HCBD February 2003 9-13
-------
10.0 REFERENCES
Abdelghani, A.A., Y.V. Pramer, T.K. Mandal, et al. 1995. Levels of toxicities of selected inorganic
and organic contaminants in a swamp environment. J. Environ. Sci. Health. Part B: Pest., Food
Contam. and Agric. Wastes. B30(5):717-731.
Anders, M.W., A.A. Elfarra and L.H. Lash. 1987. Cellular effects of reactive intermediates:
nephrotoxicity of ^-conjugates of amino acids. Arch. Toxicol. 60:103-108.
Anders, M.W. and W. Dekant. 1998. Glutathione-dependentbioactivation of haloalkenes. Ann. Rev.
Pharmacol. Toxicol. 38:501-537.
Atkinson, R. 1987. A structure-activity relationship for the estimation of rate constants for the gas-
phase reactions of OH radicals with organic compounds. Int. J. Chem. Kinet. 19:799-828 (as cited
in HSDB, 2000).
Atkinson, R. And W.P.L. Carter. 1984. Kinetics and mechanisms of the gas-phase reaction of ozone
with organic compounds under atmospheric conditions. Chem. Rev. 84: 437-70 (as cited in HSDB,
2000).
ATSDR. 1993. Supplemental Document for Hexachlorobutadiene. Agency for Toxic Substances and
Disease Registry, U.S. Dept. Health and Human Services, Public Health Service, Atlanta, GA.
ATSDR. 1994. Toxicological Profile for Hexachlorobutadiene. TP-93/08. Agency for Toxic
Substances and Disease Registry. U.S. Department of Health and Human Services, Public Health
Service. 135 pp. + Appendices.
ATSDR. 1995. ToxFAQ for Hexachlorobutadiene. Agency for Toxic Substances and Disease
Registry, Atlanta, GA. Available on the Internet at http://www.atsdr.cdc.gov/tfacts42.html Last
modified September, 1995.
ATSDR. 2000. Hazardous Substance Release and Health Effects Database. Agency for Toxic
Substances and Disease Registry, Atlanta, GA. Available on the Internet at:
http://www.atsdr.cdc.gov/hazdat.htm. Last modified August 19, 2000.
Bach, P.H., C.P Ketley and I. Ahmed. 1986. The mechanisms of target cell injury by nephrotoxins.
Food Chem. Toxicol. 24:775-779.
Badaeva, L.N., L.M. Ovsyannikova and N.I. Kiseleva. 1985. [Manifestation of neurotoxic effect of
chloroorganic pesticide hexachlorobutadiene during postnatal period of ontogenesis in rats]. Arch.
Anat. Gistol. Embriol. 89:44-49 (original in Russian) (as cited in WHO, 1994.)
Badaeva, L.N. 1983. Structural and metabolic indexes of the postnatal neurotoxicity of
organochloride pesticides. Dopov. Akad. Nauk. Ukr. pp. 55-58. (CA 099:083463W) (original in
Russian) (as cited in U.S. EPA, 199la).
HCBD February 2003 10-1
-------
Bai, C.L., P.J. Canfield andN.H. Stacey. 1992. Effects of hexachloro-1,3-butadiene and 1,1,2,2-
tetrachloroethylene on individual serum bile acids. Toxicol. Ind. Health. 8:191-203.
Birner, G., M. Werner, M.M. Ott, et al. 1995. Sex differences in hexachlorobutadiene
biotransformation and nephrotoxicity. Toxicol. Appl. Pharmacol. 132:203-212
Birner, G., M. Werner, E. Rosner, et al. 1998. Biotransformation, excretion, and nephrotoxicity of
the hexachlorobutadiene metabolite (£)-jV-acetyl-,S'-(l,2,3,4,4-pentachlorobutadienyl)-L-cysteine
sulfoxide. Chem. Res. Toxicol. 11:750-757.
Birner, G., U. Bernauer and M. Werner. 1997. Biotransformation, excretion and nephrotoxicity of
haloalkene-derived cysteine-^-conjugates. Arch. Toxicol. 72:1-8
Bogen, K.T. 1989. Cell Proliferation Kinetics and Multistage Cancer Risk Models. J. Natl. Cane.
Inst. 81:267-277.
Borst, P. andL.A. Grivell. 1978. The mitochondrial genome of yeast. Cell 15: 705-723.
Bristol, D.W., H.L. Crist, R.G. Lewis, etal. 1982. Chemical analysis of human blood for assessment
of environmental exposure to semivolitile organochlorine chemical contaminants. J. Anal. Toxicol.
6:269-275.
Burkatskaya, E.N., V.F. Viter, Z.V. Ivanova, et al. 1982. [Clinico-hygienic data on working
conditions during use of hexachlorobutadiene in vineyards]. Vrach. Delo. 11:99-102 (original in
Russian) (as cited in WHO, 1994).
Cadmus. 2000. Methods for Estimating Contaminant Occurrence and Exposure in Public Drinking
Water Systems in Support of CCL Determinations. Draft report submitted to EPA for review July
25, 2000.
Cadmus. 2001. Occurrence estimation methodology and occurrence findings report for six-year
regulatory review. Draft report to U.S. EPA, Washington, D.C., by Cadmus Group, Waltham, MA,
Octobers, 2001.
Camanzo, J., C.P. Rice and DJ. Jude. 1987. Organic priority pollutants in near-shore fish from 14
Lake Michigan tributaries and embayments, 1983. J. Great Lakes Res. 13:296-309 (as cited in
ATSDR, 1994).
ChemlDplus. 2000. Division of Specialized Information Services, National Library of Medicine
(NLM). http://chem.sis.nlm.nih.gov/chemidplus/
Chen, W., A.T. Khan, G. Fu, et al. 1999. Adsorption-desorption behaviors of hydrophobic organic
compounds in sediments of Lake Charles, Louisana, USA. Environ. Toxicol. Chem. 18(8):1610-
1616.
HCBD February 2003 10-2
-------
Chen, J.C., J.L. Stevens, A.L. Trifilis, et al. 1990. Renal cysteine conjugate betalyase-mediated
toxicity studied with primary cultures of human proximal tubular cells. Toxicol. Appl. Pharmacol.
103:463-473.
Chudin, V.A., Z.A. Gafieva, N.A. Koshurnikova, et al. 1985. Evaluating the mutagenicity and
carcinogenicity of hexachlorobutadiene. Gig. Sanit. pp. 79-80. (CA 103:033344P) (as cited in U.S.
EPA, 199 la).
Class, T. and U.K. Ballschmiter. 1987. Global baseline pollution studies: X. Atmospheric
halocarbons: global budget estimations for tetrachloroethene, 1,2-dichloroethane, 1,1,1,2-tetra
chloroethane, hexachloroethane, and hexachlorobutadiene. Estimation of the hydroxyl radical
concentrations in the troposphere of the northern and southern hemisphere. Frenseniuos' Anal.
Chem. 327(2): 198-204 (as cited in U.S. EPA, 1999).
Davis, M.E., W.O. Berndt and H.M. Mehendale. 1980. Disposition nephrotoxicity of hexachloro-
1,3-butadiene. Toxicology 16:179-191.
De Ceaurriz, J., F. Gagnaire, M. Ban, et al. 1988. Assessment of the relative hazard involved with
airborne irritants with additional hepatotoxic or nephrotoxic properties in mice. J. Appl. Toxicol.
8:417-422.
Dekant, W. 1996. Biotransformation and renal processing of nephrotoxic agents. Arch. Toxicol.
18(Suppl.): 163-172.
Dekant, W. and S. Vamvakas. 1993. Glutathione-dependent bioactivation of xenobiotics.
Xenobiotica 23:873-887.
Dekant, W., S. Vamvakas, K. Berthold, et al. 1986. Bacterial p-lyase mediated cleavage and
mutagenicity of cysteine conjugates derived from the nephrocarcinogenic alkenes trichloroethylene,
tetrachloroethylene and hexachlorobutadiene. Chem. Biol. Interact. 60:31-45.
Dekant, W., D. Schrenk, S. Vamvakas, et al. 1988a. Metabolism of hexachloro-1,3-butadiene in
mice: in vivo and in vitro evidence for activation by glutathione conjugation. Xenobiotica 18:803-
816.
Dekant, W., S. Vamvakas, K. Berthold, et al. 1988b. Enzymatic conjugation of hexachloro-1,3-
butadiene with glutathione. Formation of l^glutathion-^-y^-l^S^^-pentachlorobuta-l^-diene
and l,4-bis(glutathion-,S'-yl)-l,2,3,4-tetrachlorobuta-l,3-diene. DrugMetab. Dispos. 16:701-716.
Dekant, W., S. Vamvakas, M. Koob, et al. 1990. A mechanism of haloalkene-induced renal
carcinogenesis. Environ. Health Perspect. 88:107-110.
Dekant, W., G. Urban, C. Gorsman, et al. 1991. Thioketene formation from y-haloalkene 2-
nitrophenyl disulfides: models for biological reactive intermediates of cytotoxic ^-conjugates. J. Am.
Chem. Soc. 113:5120-5122.
HCBD February 2003 10-3
-------
Dekant, W., G. Birner, M. Werner, et al. 1998. Glutathione conjugation of perchloroethene in
subcellular fractions from rodent and human liver and kidney. Chem. Biol. Interact. 116:31-43.
DeMeester, C., M. Mercier and F. Poncele. 1981. Mutagenic activity of butadiene,
hexachlorobutadiene, and isoprene. Ind. Environ. Xenobio. Proc. Intl. Conf. pp. 195-203.
DeVault, D.S. 1985. Contaminants in fish from Great Lakes harbors and tributary mouths. Arch.
Environ. Contam. Toxicol. 14:578-597.
DiNovi, M. 1997. FDA, Chemistry Review Branch, Office of Premarket Approval. Personal
communication with Denis Borum, U.S. Environmental Protection Agency, Office of Water,
Planning and Standards. August 21 and 22 (as cited in U.S. EPA, 1999).
Driscoll, T.R., H.H. Hamdan, G. Wang, et al. 1992. Concentrations of individual serum or plasma
bile acids in workers exposed to chlorinated aliphatic hydrocarbons. Br. J. Ind. Med. 49:700-705.
Duprat, P. andD. Gradiski. 1978. Percutaneous toxicity of hexachlorobutadiene. Acta. Pharmacol.
Toxicol. 43:346-353.
Gage, J.C. 1970. The subacute inhalation toxicity of 109 industrial chemicals. Br. J. Ind Med. 27:1-
18.
Galloway, S.M., MJ. Armstrong, C. Reuben, et al. 1987. Chromosome aberrations and sister
chromatid exchanges in Chinese hamster ovary cells: evaluations of 108 chemicals. Environ. Mol.
Mutagen. 10(Suppl.):l-175.
Garle, M. and J. Fry. 1989. Detection of reactive metabolites in vitro. Toxicology 54:101-110.
Gehring, PJ. and D. MacDougall. 1971. Review of the toxicity of hexachlorobenzene and
hexachlorobutadiene. Dow Chemical Co., Midland, MI. (as cited in NTP, 1991).
German, IV. 1986. [Level of chromosome aberrations in workers coming in contact with
hexachlorobutadiene during production.] Gig. Tr. Prof. Zabol. 5:57-79. (original in Russian) (as
cited in WHO, 1994).
German, IV. 1988. [On mutagenic activity of the pesticide hexachlorobutadiene.] Citol Gen. 22:40-
42. (original in Russian) (as cited in WHO, 1994).
Gess, P. and S.G. Pavlostathis. 1997. Desorption of chlorinated organic compounds from a
contaminated estuarine sediment. Environ. Toxicol. Chem. 16(8): 1598-1605.
Gietl, Y.S. and M.W. Anders. 1991. Biosynthesis and biliary excretion of S-conjugates of
hexachlorobuta-l,3-diene in the perfused rat liver. Drug Metab. Dispos. 19:274-277.
Gietl, Y. and M.W. Anders. 1990. Formation and excretion of the glutathione ^-conjugate of
hexachlorobutadiene in the perfused rat liver. Toxicologist 10:199.
HCBD February 2003 10-4
-------
Gietl, Y., S. Vamvakas andM.W. Anders. 1991. Intestinal absorption of S-pentachlorobutadienyl-
glutathione and ^-(pentachlorobutadieny^-L-cysteine, the glutathione and cysteine S-conjugates of
hexachlorobuta-l,3-diene. DrugMetab. Dispos. 19:703-707.
Gilliom, R.J., D.K. Mueller, and L.H. Nowell. In press. Methods for comparing water-quality
conditions among National Water-Quality Assessment Study Units, 1992-95. U.S. Geological
Survey Open-File Report 97-589.
Gradiski, D., P. Duprat, J.L. Magadur, et al. 1975. [Toxicological and experimental study of
hexachlorobutadiene]. Eur. J. Toxicol. 8:180-187 (original in French) (as cited in U.S. EPA, 199 la).
Green, T. and J. Odum. 1985. Structure/activity studies of the nephrotoxic and mutagenic action of
cysteine conjugates of chloro- and fluoroalkenes. Chem. Biol. Interact. 54:15-31.
Green, T., J. Odum, J.A. Nash, et al. 1990. Perchloroethylene-induced rat kidney tumors: An
investigation of the mechanisms involved and their relevance to humans. Toxicol. Appl. Pharmacol.
103:80-89.
Grosjean, E. and R. A. Rassmussen. 1999. Toxic air contaminants in Porto Alegre, Brazil. Environ.
Sci. Tech. 33:1970-1978.
Groves, C.E., R.G. Schnellmann, P.P. Sokol, etal. 1991. Pentachlorobutadienyl-L-cysteine (PCBC)
toxicity: the importance of mitochondrial dysfunction. J. Biochem. Toxicol. 6:253-260.
Hardin, B.D., G.P. Bond, M.R. Sikov, et al. 1981. Testing of selected workplace chemicals for
teratogenic potential. Scand. J. Work Environ. Health 7(Suppl. 4):66-75.
Harleman, J.H. and W. Seinen. 1979. Short-term toxicity and reproduction studies in rats with
hexachloro-(l,3)-butadiene. Toxicol. Appl. Pharmacol. 47:1-14.
Harris, S.J., G.P. Bond and R.W. Niemeier. 1979. The effects of 2-nitropropane, naphthalene, and
hexachlorobutadiene on fetal rat development [abstract]. Toxicol. Appl. Pharmacol. 43:A35.
Haworth, S., T. Lawlor, K. Mortelmans, et al. 1983. Salmonella mutagenicity test results for 250
chemicals. Environ. Mutagen. l(Suppl.):3-142.
Hendricks, A. J., H. Pieters and J. De Boer. 1998. Accumulation of metals, polycyclic (halogenated)
aromatic hydrocarbons, and biocides in zebra mussel and eel from the Rhine and Muese Rivers.
Environ.Toxicol. Chem. 17(10): 1885-1898.
Hinchman, C.A. and N. Ballatori. 1990. Glutathione-degrading capacities of liver and kidney in
different species. Biochem. Pharmacol. 40:1131-1135.
Hook, J.B., J. Ishmael and E.A. Lock. 1983. Nephrotoxicity of hexachloro-l,3-butadiene in the rat:
the effect of age, sex, and strain. Toxicol. Appl. Pharmacol. 67:122-131.
HCBD February 2003 10-5
-------
Hook, J.B.,M.S. Rose and E. A. Lock. 1982. The nephrotoxicity of hexachloro-1,3-butadiene in the
rat: studies of organic anion and cation transport in renal slices and the effect of monooxygenase
inducers. Toxicol. Appl. Pharmacol. 65:373-382.
Howard, P.H. 1989. Handbook of Environmental Fate and Exposure Data for Organic Chemicals.
Volume 1 Large Production and Priority Pollutants. Chelsea, MI: Lewis Publishers, Inc. 574 pp.
HSDB. 2000. Hazardous Substance Data Bank. Hexachlorobutadiene. Division of Specialized
Information Services, National Library of Medicine, http://toxnet.nlm.nih.gov/. Last revised
03/30/2000.
IARC. 1979. Monographs on the evaluation of the carcinogenic risk of chemicals to humans.
International Agency for Research on Cancer 20:179-193 (as cited in U.S. EPA, 1999).
IARC. 1999. Monographs on the evaluation of the carcinogenic risk of chemicals to humans.
International Agency for Research on Cancer 73:277-294.
Ishmael J, Lock EA. 1986. Nephrotoxi city of hexachlorobutadiene and its glutathi one-derived
conjugates. In: Presented at the Fourth International Symposium of the Society of Toxicologic
Pathologists, June 5-7, 1985. Washington, DC. Toxicol Pathol 14:258-262.
Ishmael J, Pratt I, Lock EA. 1982. Necrosis of the pars recta (S, segment) of the rat kidney produced
by hexachloro-1,3-butadiene. J Pathol 138:99-113.
Ishmael J, Pratt I, Lock EA. 1984. Hexachl oro-1,3-butadiene-induced renal tubular necrosis in the
mouse. J Pathol 142:195-203.
Jaffe, D.R., C.D. Hassall and K. Brendel. 1983. In vivo and in vitro nephrotoxi city of the cysteine
conjugate of hexachlorobutadiene. J. Toxicol. Environ. Health 11:857-867.
Johnson, L.D. and J.C. Young. 1983. Inhibition of anaerobic digestion by organic priority pollutant
chemicals. J. Water Pollut. Control Fed. 55:1441-1449 (as cited in HSDB, 2000).
Jones, T.W., C. Quin, V.H. Schaeffer, et al. 1988. Immunohistochemical localization of glutamine
transaminase K, a rat kidney cysteine conjugate p-lyase, and the relationship to the segment
specificity of cysteine conjugate specificity. Mol. Pharmacol. 34:621-627.
Jones, T.W., A. Wallin, H. Thor, et al. 1986. The mechanism of pentachlorobutadienyl-glutathione
nephrotoxi city studied with isolated rat epithelial cells. Arch. Biochem. Biophys. 251:504-513.
Jones, T.W., R.G. Gerdes, K. Ormstad, et al. 1985. The formation of both a mono- and bis-
substituted glutathione conjugate of hexachlorobutadiene by isolated hepatocytes and following in
vivo administration to the rat. Chem. Biol. Interact. 56:251-267'.
HCBD February 2003 10-6
-------
Jonker, D., M.A. Jones, PJ. van Bladeren, et al. 1993a. Acute (24 hr) toxicity of a combination of
four nephrotoxicants in rats compared with the toxicity of the individual compounds. Food Chem.
Toxicol. 31:45-52.
Jonker, D., R.A. Woutersen, P.J. van Bladeren, et al. 1993b. Subacute (4-wk) oral toxicity of a
combination of four nephrotoxins in rats: comparison with the toxicity of the individual compounds.
Food Chem. Toxicol. 31:125-136.
Kennedy GL, Graepel GJ. 1991. Acute toxicity in the rat following either oral or inhalation
exposure. Toxicol Lett 56:317-326.
Kim, H.S., S.H. Cha, D.G. Abraham, et al. 1997. Intranephron distribution of cysteine-^-conjugate
P4yase activity and its implications for hexachloro-l,3-butadiene-induced nephrotoxi city in rats.
Arch. Toxicol. 71:131-141.
Kirby, G.M. and P.H. Bach. 1995. Enhanced hexachloro-1:3-butadiene nephrotoxi city in rats with
a preexisting adriamycin-induced nephrotic syndrome. Toxicol. Pathol. 23:303-312.
Kociba, R.J., P.J. Gehring and C.G. Humiston. 1971. Toxicologic study of female rats administered
hexachlorobutadiene or hexachlorobenzene for thirty days. Dow Chemical Company. Chemical
Biological Research, Midland, MI. (unpublished report) (as cited in U.S. EPA, 1991a).
Kociba, R.J., D.G. Keyes, G.C. Jersey, et al. 1977. Results of a 2-year chronic toxicity study with
hexachlorobutadiene in rats. Am. Ind. Hyg. Assoc. J. 38:589-602.
Kolpin, D.W., J.E.BarbashandR.J. Gilliom. 1998. Occurrence of pesticides in shallow groundwater
of the United States: initial results from the National Water Quality Assessment Program. Environ.
Sci. Technol. 32:558-566.
Koob, M. and W. Dekant. 1992. Biotransformation of the hexachlorobutadiene metabolites 1-
(glutathion-S-yO-pentachlorobutadiene and l^cystein-S-y^-pentachlorobutadiene in the isolated
perfused rat liver. Xenobiotica 22:125-138.
Krasniuk, E.P., L.A. Ziritskaya, V.G. Bioko, et al. 1969. [Health conditions of vine-growers
contacting with fumigants hexachorobutadiene and polychlorbutan-80.] Vrach. Delo.7:lll-115
(original in Russian) (as cited in U.S. EPA, 1991a).
Kuehl, D.W., B. Butterworth and P.J. Marquis. 1994. A national study of chemical residues in fish.
III. Study results. Chemosphere 29(3):523-535.
Kuo, C.H. and J.B. Hook. 1983. Effects of age and sex on hexachloro-1,3-butadiene toxicity in the
Fischer 344 rat. Life Sci. 33:517-523.
Kusznesof, P. 1997. FDA Office of Premarket Approval. Personal communication with Amy
Benson. Abt. Assoc. Inc. May 28, 1997 (as cited in U.S. EPA, 1999).
HCBD February 2003 10-7
-------
Lapham, W.W., K.M. Neitzert, M. J. Moran, et al. 1997. USGS compiles data set for national
assessment of VOCs in ground water. Ground Water Monit. Remed. 17(4):147-157.
Larson, S.J., RJ. Gilliom and P.D. Capel. 1999. Pesticides in Streams of the United StatesInitial
Results from the National Water Quality Assessment Program. U.S. Geological Survey
Water-Resources Investigations Report 98-4222. 92 pp. Available on the Internet at:
http://water.wr.usgs.gov/pnsp/rep/wrir984222
Lash, L.H., A.A. Elfarra and M.W. Anders. 1986. Renal cysteine conjugate p-lyase: Bioactivation
of nephrotoxic cysteine ^-conjugates in mitochondrial outer membrane. J. Biol. Chem 261:5930-
5935.
Lash, L.H., R.M. Nelson, R.A. Dyke, et al. 1990. Purification and characterization of human kidney
cytosolic conjugate beta-lyase activity. Drug Metab. Dispos. 18:50-54.
Laska, A.L., C.K. Bartell and J.L. Laseter. 1976. Distribution of hexachlorobenzene and
hexachlorobutadiene in water, soil, and selected aquatic organisms along the lower Mississippi
River, Louisana. Bull. Environ. Contam. Toxicol. 15:535-542.
Leahy, P.P. and T.H. Thompson. 1994. The National Water-Quality Assessment Program. U.S.
Geological Survey Open-File Report 94-70. 4 pp. Available on the Internet at:
http://water.usgs.gov/nawqa/NAWQA.OFR94-70.html Last updated August 23, 2000.
Leeuwangh, P., et al. 1975. Toxicity of hexachlorobutadiene in aquatic organisms: sublethal effects
of toxic chemicals on aquatic animals. In: Proceedings of the Swedish-Netherlands Symposium,
September 2-5, New York, NY. Elsevier Scientific Publishing Co., Inc. (as cited in U.S. EPA,
199 la).
Levins, P., J. Adams, P. Brenner, et al. 1979. Sources of toxic pollutants found in influents to
sewage treatment plants VI. Integrated interpresentation. NTIS PB81-219685 (as cited in ATSDR,
1994).
Lock, E.A. 1988. Studies on the mechanism of nephrotoxicity and nephrocarcinogenicity of
halogenated alkenes. CRC Crit. Rev. Toxicol. 19:23-42.
Lock, E.A. 1994. The role of mechanistic studies in understanding target organ toxicity. Arch.
Toxicol. 16(Suppl.):151-160.
Lock, E.A., J. Ishmael and J.B. Hook. 1984. Nephrotoxicity of hexachloro-1,3-butadiene in the
mouse: The effect of age, sex, strain, monooxygenase modifiers, and the role of glutathione. Toxicol.
Appl. Pharmacol. 72:484-494.
Lock, E.A., Y. Sani, R.B. Moore, et al. 1996. Bone marrow and renal injury associated with
haloalkene cysteine conjugates in calves. Arch. Toxicol. 70:607-619.
HCBD February 2003 10-8
-------
Lock, E.A. and J. Ishmael. 1979. The acute toxic effects of hexachloro-1,3-butadiene on the rat
kidney. Arch. Toxicol. 19:23-42.
Lopes, T. J. and S.G. Dionne. 1998. A Review of Semivolatile and Volatile Organic Compounds in
Highway Runoff and Urban Stormwater. U.S. Geological Survey Open-File Report 98-409. 67 pp.
MacFarlane, M., J.R. Foster, G.G. Gibson, et al. 1989. Cysteine conjugate beta-lyase of rat kidney
cytosol: characterization, immunocytochemical localization, and correlation with
hexachlorobutadiene nephrotoxicity. Toxicol. Appl. Pharmacol. 98:185-197.
Mansouri, A., B. Fromenty, A. Benson et al., 1997. Multiple hepatic mitochondrial DNA deletions
suggest premature oxidative aging in alcoholic patients. Journal of Hepatology 27: 96-102.
McConnell, G., D.M. Ferguson and C.R. Pearson. 1975. Chlorinated hydrocarbons and the
environment. Endeavor 34:13-18.
McLellan, L.I., C.R. Wolf and J.D. Hayes. 1989. Human microsomal glutathione S-transferase: Its
involvement in the conjugation of hexachlorobuta-l,3-diene with glutathione. J. Biochem. 258:87-
93.
Mes, J., D. J. Davies and D. Turton. 1985. Environmental contaminants in human fat: A comparison
between accidental and nonaccidental causes of death. Ecotoxicol. Environ. Safety 10(l):70-74.
Morel, G., M. Ban, D. Hettich, et al. 1999. Role of SAM-dependant thiolmethylation in the renal
toxicity of several solvents in mice. J. Appl. Toxicol. 19(l):47-54.
Nakagawa, Y., Y. Kitahori, M. Cho, et al. 1998. Effects of hexachloro-l,3-butadiene on renal
carcinogenesis in male rats pretreated with 7V-ethyl-7V-hydroxyethylnitrosamine. Toxicol. Pathol.
26:361-366.
Nash, J.A., LJ. King, E.A. Lock, et al. 1984. The metabolism and disposition of hexachloro-1,3-
butadiene in the rat and its relevance for nephrotoxicity. Toxicol. Appl. Pharmacol. 73:124-137.
NIOSH. 1981. Tierllmutagenic screening of 13 NIOSHpriority compounds: Individual compound
report hexachloro-1,3-butadiene. National Institute on Occupational Safety and Health, Cincinnati,
OH.
NTP. 1991. Toxicity studies of hexachloro-1,3-butadiene in B6C3Fj mice (feed studies). National
Toxicology Program U.S. Department of Health and Human Services, Public Health Service,
National Institute of Health, Research Triangle Park, NC. NIH Publication No. 91-3120.
Oesch, F. and C.R. Wolf. 1989. Properties of the microsomal and cytosolic glutathione transferases
involved in hexachloro-1,3-butadiene conjugation. Biochem. Pharmacol. 38:353-359.
Olea, N., F. Olea-Serrano, P. Lardelli-Claret, et al. 1999. Inadvertent exposure to xenoestrogens in
children. Toxicol. Indust. Health. 15(1-2):151-158.
HCBD February 2003 10-9
-------
Oliver, E.G. and R.A. Bourbonniere. 1985. No title given. J. Great Lakes Res. 11:366 (as cited in
Choudhary, 1995).
Oliver, E.G. and A. J. Nimi. 1983. Bioconcentrations of chlorobenzenes from water by rainbow trout:
correlations with partition coefficients and environmental residues. Environ. Sci. Tech. 10:148-152
(as cited in AT SDR, 1994).
OSHA. 1989. Occupational Safety and Health Administration: Part III. Federal Register
54:2332-2959. January 19.
Pahler, A., G. Birner, M.M. Ott, et al. 1997. Binding of hexachlorobutadiene to a2(J-globulin and its
role in nephrotoxicity in rats. Toxicol. Appl. Pharmacol. 147:372-380.
Pahler, A., K. Blumbach, J. Herbst, et al. 1999. Quantitation of a2(J-globin in rat kidney cytosol by
capillary electrophoresis. Anal. Biochem. 267:203-211.
Payan, J.P., D. Beydon, J.P. Fabry, et al. 1993. Partial contribution of biliary metabolites to
nephrotoxicity, renal content, and excretion of [14C]hexachloro-1,3-butadiene in rats. J. Appl.
Toxicol. 13:19-24.
Payan, J.P., J.P. Fabry, D. Beydon, et al. 1991. Biliary excretion of hexachloro-1,3-butadiene and
its relevance to tissue uptake and renal excretion in male rats. J. Appl. Toxicol 11:437-442.
Pellizari, E.D., M.D. Erickson andR.A. Zweidinger. 1979. Formulation of a preliminary assessment
of halogenated organic compounds in man and environmental media. EPA 560/13-179-00. Research
Triangle Inst, Research Triangle Park, NC.
Pellizari, E.D. 1978. Quantification of chlorinated hydrocarbons in previously collected air samples.
EPA-450/3-78-112.
Pellizari, E.D. 1982. Analysis for organic vapor emissions near industrial and chemical waste
disposal sites. Environ. Sci. Tech. 16:781-785.
Perry, S., H. Harries, C. Scholfield, et al. 1995. Molecular cloning and expression of cDNA for
human kidney cysteine conjugate p-lyase. FEES Letters 360:227-280.
Prytula, M.T. and S.G. Pavlostathis. 1996. Effect of contaminant and organic matter bioavailability
onthemicrobial dehalogenation of sediment-bound chlorobenzenes. Water Res. 30(11):2669-2680.
Rapson, W., M.A. Nazar and V.V. Busky. 1980. Mutagenicity produced by aqueous chlorination
of organic compounds. Bull. Environ. Contam. Toxicol. 24:590-596.
Reichert, D., T. Neudecker and U. Spengler. 1983. Mutagenicity of dichloroacetylene and its
degradation products trichloroacetyl chloride, trichloroacryloyl chloride and hexachlorobutadiene.
Mutat. Res. 117:21-29.
HCBD February 2003 10-10
-------
Reichert, D., S. Schutz and M. Metzler. 1985. Excretion pattern and metabolism of
hexachlorobutadiene in the rat: Evidence for metabolic activation by conjugation reactions.
Biochem. Pharmacol. 34:499-505.
Reichert, D. and S. Schutz. 1986. Mercapturic acid formation is an activation and intermediary step
in the metabolism of hexachlorobutadiene. Biochem. Pharmacol. 35:1271-1275.
Reichert, D. 1983. Metabolism and disposition of hexachloro(l,2)butadienein rats. In: Hayes, A.W.,
Schnell, R.C., Miya, T.S. (Eds.), Developments in the Science and Practice of Toxicology. Elsevier
Publishers, pp.411-414.
Reichert, D., T. Neudecker and S. Schutz. 1984. Mutagenicity of hexachlorobutadiene, perchloro-
butenoic acid and perchlorobutenoic acid chloride. Mutat. Res. 137:89-93.
Roldan-Arjona, T., M. Garcia-Pedrajas and F. Luque-Romero. 1991. An association between
mutagenicity of the Ara test of Salmonella typhimurium and carcinogenicity in rodents for 16
halogenated aliphatic hydrocarbons. Mutagenesis 6:199-205.
Rosner, E., M. Miiller and W. Dekant. 1998. Stereo- and regioselective conjugation of ^-halovinyl
mercapturic acid sulfoxides by glutathione ^-transferases. Chem. Res. Toxicol. 11:12-18.
Saillenfait, A.M., P. Bonnet, J.P. Guenier, et al. 1989. Inhalation teratology study onhexachloro-1,3-
butadiene in rats. Toxicol. Lett. 47:235-240.
Saito, K., S. Uwagawa, H. Kaneko, et al. 1996. a2(J-globulins in the urine of male rats: A reliable
indicator for a2(J-globulin accumulation in the kidney. Toxicology 106:149-157.
Schiffman, D., D. Reichert and D. Henschler. 1984. Induction of morphological transformation and
unscheduled DNA synthesis in Syrian hamster embryo fibroblasts by hexachlorobutadiene and its
putative metabolite pentachlorobutenoic acid. Cancer Lett. 23:297-305.
Schnellmann, R., E.A. Lock and L. Mandel. 1987. A mechanism of S-(l,2,3,4,4-pQntach\oro-l,3-
butadienyl)-L-cysteine toxicity to rabbit renal proximal tubules. Toxicol. Appl. Pharmacol. 90:513-
521.
Schrenk, D. and W. Dekant. 1989. Covalent binding of hexachlorobutadiene metabolites to renal
and hepatic DNA. Carcinogenesis 10:1139-1141.
Schwetz, B.A., F.A. Smith and C.G. Humiston. 1977. Results of a reproduction study in rats fed
diets containing hexachlorobutadiene. Toxicol. Appl. Pharmacol. 42:387-398.
Shah, JJ. andE.K. Heyerdahl. 1988. National Ambient VOCs Database Update. Report by Nero and
Assoc. Inc. Portland, OR. to U.S. Environmental Protection Agency., Atmos. Sci. Res. Lab.,
Research Triangle Park, NC. EPA600/3-88/010a.
HCBD February 2003 10-11
-------
Shaw, L.M., J.W. London and L.E. Petersen. 1978. Isolation of y-glutamyltransferase from human
liver, and comparison with the enzyme from human kidney. Clin. Chem. 24:905-915.
Simmon, V.F. 1977. Structural correlation of carcinogenic and mutagenic alkyl halides. In: Proc.
2nd FDA Office of Science Summer Symposium, U.S. Naval Academy, Aug. 31-Sept. 2, 1977. pp.
163- 171 (as cited in Stott et al., 1981).
Singh, H.B., LJ. Sales, A. Smith, et al. 1980. Atmospheric measurements of selected hazardous
organic chemicals. Menlo Park, CA: SRI Inter. Project No. 7774:6 (as cited in ATSDR, 1994).
Singh, H.B., L. J. Sales and R.E. Stiles. 1982. Distribution of selected gaseous organic mutagens and
suspect carcinogens in ambient air. Environ. Sci. Tech. 16:872-880 (as cited in ATSDR, 1994).
Spicer, C.W., B. Buxton, M.W. Holdren, et al. 1996. Variability of hazardous air pollutants in an
urban area. Atmospher. Environ. 30(20):3443-3456.
Squillace, P. J., MJ. Moran, W.W. Lapham, et al. 1999. Volatile organic compounds in untreated
ambient groundwater of the United States, 1985-1995. Environ. Sci. Technol. 33(23):4176-4187.
Staples, C.A., A.F. Werner and TJ. Hoogheem. 1985. Assessment of priority pollutant
concentrations in the United States using Storet databases. Environ. Toxicol. Chem. 4:131-142.
Stevens, J.L., J.D. Robbins and R.A. Byrd. 1986. A purified cysteine conjugate P-lyase from rat
kidney cytosol. J. Biol. Chem. 261:15529-15537.
Stevens, J.L. 1985. Cysteine conjugate P-lyase activities in rat kidney cortex: Subcellular
localization and relationship to the hepatic enzyme. Biochem. Biophys. Res. Commun. 129:499-504.
Stott, W.T., J.F. Quast and P.G. Watanabe. 1981. Differentiation of the mechanisms of oncogenicity
of 1,4-dioxane and 1,3-hexachlorobutadiene in the rat. Toxicol. Appl. Pharmacol. 60:287-300.
Tabak, H.H., S.A. Quave, C.E. Mashni, et al. 1981. Biodegradability studies with organic priority
pollutant compounds. J. Water Pollut. Control Fed. 53:1503-1518.
Tchounwou, P.B., A.A. Abdelghani, Y.V. Pramar, et al. 1998. Health risk assessment of
hexachlorobenzene and hexachlorobutadiene residues in fish collected from a hazardous waste
contaminated wetland in Louisiana, USA. Proceedings of the 1997 7th Symposium on Toxicol. And
Risk Assess: Ultrviolet Radiation and the Environ. Conshoshocken, PA: American Society for
Testing and Materials. ASTM Special Pub. V 1333.
Theiss, J.C., G.D. Stoner, M.B. Shimkin, et al. 1977. Test for carcinogenicity of organic
contaminants of United States drinking waters by pulmonary tumor response in strain A mice.
Cancer Res. 37:2717-2720.
Trevisan, M., E. Graviani, A.M.A. Del Re, et al. 1998. Formation of jV-nitrosoterbuthylazine and
7V-nitrosoterbutryn in a model system of soil water. J. Agricul. Food. Chem. 46(1):314-317.
HCBD February 2003 10-12
-------
U.S. DOE. 2001. RCRA Corrective Measures Study Plan for the Lawrence Berkeley National
Laboratory. United States Department of Energy.
U.S. EPA. 1976. An Ecological Study of Hexachlorobutadiene. U.S. Environmental Protection
Agency, Office of Toxic Substances, Washington D.C. EPA 560/6-76-010.
U.S. EPA. 1980. Ambient Water Quality Criteria for Hexachlorobutadiene. U.S. Environmental
Protection Agency, Criteria and Standards Division, Washington D.C. EPA 44/5-80-053; PB-81-
117640. (as cited in ATSDR, 1994; WHO, 1994; HSDB, 2000).
U.S. EPA. 1984. Health Effects Assessment for Hexachlorobutadiene. U.S. Environmental
Protection Agency, Office of Emergency and Remedial Response, Office of Solid Waste and
Emergency Response, Washington, D.C. EPA/540/1-86-053.
U.S. EPA. 1986a. Guidelines for Carcinogen Risk Assessment. United States Environmental
Protection Agency. Federal Register 51(185):33992-34003.
U.S. EPA. 1986b. Guidelines for the Health Risk Assessment of Chemical Mixtures. U. S.
Environmental Protection Agency. Federal Register 51(185):34014-34025.
U.S. EPA. 1986c. Guidelines for Mutagenicity Risk Assessment. U.S. Environmental Protection
Agency. Federal Register 51(185):34006-34012.
U.S. EPA. 1987. National Primary Drinking Water Regulations-Synthetic Organic Chemicals;
Monitoring for Unregulated Contaminants; Final Rule. U.S. Environmental Protection Agency. July
8. Federal Register, vol. 52, no. 130, 25720 [52 FR 25720].
U.S. EPA. 1988. Recommendations for and Documentation of Biological Values for Use in Risk
Assessment. United States Environmental Protection Agency, Environmental Criteria and
Assessment Office, Office of Health and Environmental Assessment, Cincinnati, Ohio. EPA-600/6-
87-008.
U.S. EPA. 1990. United States Environmental Protection Agency. Part III. Federal Register. 503:
47229.
U.S. EPA. 1991a. Drinking Water Health Advisories: Hexachlorobutadiene. In: Volatile Organic
Compounds. United States Environmental Protection Agency, Office of Drinking Water. Ann
Arbor, MI: Lewis Publishers, pp.51-68.
U.S. EPA. 1991b. Guidelines for Developmental Toxicity Risk Assessment. U.S. Environmental
Protection Agency. Federal Register 56:63798-63826.
U.S. EPA. 199 Ic. National Primary Drinking Water Regulations- Synthetic Organic Chemical sand
inorganic Chemicals; Monitoring for Unregulated Contaminants; National Primary Drinking Water
Regulations Implementation; National Secondary Drinking Water Regulations; Final Rule.
HCBD February 2003 10-13
-------
U.S. EPA. 1991d. U.S. Environmental Protection Agency. Federal Register 56:3526-3597.
U.S. EPA. 1991e. Alpha2u-globulin: Association with chemically induced renal toxicity and
neoplasia in the male rat, Risk Assessment Forum, Washington, DC. EPA/625/3-91/019F.
U.S. EPA. 1992a. National study of chemical residues in fish: Vol. I and II. United States
Environmental Protection Agency, Office of Sci. and Tech, Washington, DC. EPA 823-R-92-008a
and EPA 823-R-92-008b.
U.S. EPA. 1992b. Initial Submission: Detection of hexachlorobenzene and hexachlorobutadiene in
sediment, fish, ducks, freshwater clams, squirrels, and raccoons. 12/29/1992. United States
Environmental Protection Agency, Office of Toxic Substances. Doc. 88-930000010.
U.S. EPA. 1992c. Drinking Water; National Primary Drinking Water Regulations - Synthetic
Organic Chemicals and Inorganic Chemicals; National Primary Drinking Water Regulations
Implementation. U.S. Environmental Protection Agency. July 17. Federal Register, vol. 57,no. 138,
31776-31849 pp. [57 FR 31776]
U.S. EPA. 1992d. A cross-species scaling factor for carcinogen risk assessment based on
equivalence of mg/kg3/4 day; notice. Draft report. United States Environmental Protection Agency,
Federal Register 57(109):24152-24173.
U.S. EPA. 1994. A screening analysis of ambient monitoring data for the Urban Area Source
Program. United States Environmental Protection Agency, Office of Air Qual. Plan. And Stand.
EPA-453/R-94-075.
U.S. EPA 1995. Use of the Benchmark Dose Approach in Health Risk Assessment. U.S.
Environmental Protection Agency. EPA/630/R-94/007.
U.S. EPA. 1996a. Proposed Guidelines for Carcinogen Risk Assessment. United States
Environmental Protection Agency, Office of Research and Development, Washington, D.C.
EPA/600/P-92/003C.
U.S. EPA. 1996b. Guidelines for Reproductive Toxicity Risk Assessment. U.S. Environmental
Protection Agency, Office of Research and Development, Washington, D.C. EPA/630/R-96/009.
U.S. EPA. 1996c. Emergency Planning and Community Right-to-Know Section 313, List of Toxic
Chemicals. U.S. Environmental Protection Agency. Available on the internet at:
http://www.epa.gov/tri/chemls2.pdf. Last modified March 23, 2000. Link to site at:
http ://www. epa.gov/tri/chemical .htm
U.S. EPA. 1997a. Hexachlorobutadiene. Integrated Risk Information Service, U.S. Environmental
Protection Agency, http://www.epa.gov/iris/subst/0058.htm. Last updated 04/01/97.
U.S. EPA. 1997b. U.S. Environmental Protection Agency. Announcement of the Draft Drinking
Water Contaminant Candidate List; Notice. Fed. Reg. 62(193):52193. October 6.
HCBD February 2003 10-14
-------
U. S. EPA, 1998a. Draft Ambient Water Quality Criteria for the Protection of Human Health. Office
of Water, Washington, D.C. EPA 822-R-98-004.
U.S. EPA. 1998b. Guidelines for Neurotoxicity Risk Assessment. U. S. Environmental Protection
Agency. Federal Register 63(93):26926-26954.
U.S. EPA. 1998c. Science Policy Council Handbook: Peer Review. U.S. Environmental Protection
Agency, Office of Science Policy, Office of Research and Development, Washington, D.C.
EPA/1 OO/B-98/001.
U.S. EPA. 1998d. U.S. Environmental Protection Agency. Announcement of the Drinking Water
Contaminant Candidate List; Final Rule. Fed. Reg. 63 (274): 10273. March 2.
U.S. EPA. 1999a. Superfund Hazardous Waste Site Basic Query Form. U.S. Environmental
Protection Agency. Available on the Internet at: http://www.epa.gov/superfund/sites/query/basic.htm
Last modified December 1, 1999.
U.S. EPA. 1999b.. A Review of Contaminant Occurrence in Public Water Systems. U.S.
Environmental Protection Agency, Office of Water, Washington, D.C. EPA/816-R-99/006.
U.S. EPA. 1999c. Guidelines for Carcinogen Risk Assessment. SAB Review Draft. U.S.
Environmental Protection Agency, Office of Research and Development, Washington, D.C. NCEA-
F-0644.
U.S. EPA. 2000a. What is the Toxic Release Inventory. U.S. Environmental Protection Agency.
Available on the Internet at: http://www.epa.gov/tri/general.htm Last modified February 28, 2000.
U.S. EPA. 2000b. TRI Explorer: Trends. U.S. Environmental Protection Agency. Available on the
Internet at: http://www.epa.gov/triexplorer/trends.htm Last modified May 5, 2000.
U.S. EPA. 2000c. TRI Explorer: Are Year-to-Year Changes Comparable? U.S. Environmental
Protection Agency. Available on the Internet at: www.epa.gov/triexplorer/yearsum.htm Last
modified May 5, 2000.
U.S. EPA. 2000d. The Toxic Release Inventory (TRI) and Factors to Consider when Using TRI
Data. U.S. Environmental Protection Agency. Available on the Internet at:
http://www.epa.gov/tri/tri98/98over.pdf. Last modified August 11, 2000. Link to site at:
http ://www. epa.gov/tri/tri98
U.S. EPA. 2000e.. Water Industry Baseline Handbook, Second Edition (Draft). U.S. Environmental
Protection Agency, Washington, D.C. March 17.
U.S. EPA. 200la. Analysis of national occurrence of the 1998 Contaminant Candidate List
regulatory determination priority contaminants in public water systems. Office of Water. EPA report
815-D-01-002. 77pp.
HCBD February 2003 10-15
-------
U.S. EPA. 2001b. Occurrence of unregulated contaminants in public water systems: An initial
assessment. Office of Water. EPA report 815-P-00-001. Office of Water. 50 pp.
U.S. EPA. 2001c. Contaminant Candidate List preliminary regulatory determination support
document for heachlorobuatdiene. Office of Water. EPA report 815-R-01-009. 50 pp.
U.S. EPA. 2002. Review of the Reference Dose and Reference Concentration Processes (Draft).
Risk Assessment Forum. EPA report 630-P-02-002a.
Vamvakas, S., W. Dekant and D. Henschler. 1989. Genotoxicity of haloalkene and haloalkane
glutathione ^-conjugates in porcine kidney cells. Toxicol. In Vitro 3:151-156.
Vamvakas, S., FJ. Kordowich, W. Dekant, et al. 1988. Mutagenicity of hexachloro-1,3-butadiene
and its S conjugates in the Ames test role of activation by the mercapturic acid pathway in its
nephrocarcinogenicity. Carcinogenesis 9:907-910.
Van Duuren, B.L., B.M Goldschmidt, G. Loewengart, et al. 1979. Carcinogenicity of halogenated
olefinic and aliphatic hydrocarbons in mice. J. Natl. Cancer Inst. 63:1433-1439.
Wallace, D.C. 1999. Mitochondrial Diseases in Man and Mouse. Science 283: 1482-1488.
Wallin, A., T.W. Jones, A.E. Vercesi, et al. 1987. Toxicity of ^-pentachlorobutadienyl-L-cysteine
studied with isolated rat renal cortical mitochondria. Arch. Biochem. Biophys. 258:365-372.
Wallin A., R.G. Gerdes, R.Morgenstern, T.W. Jones, K. Ormstad. 1988. Features of microsomal and
cytosolic glutathione conjugation of hexachlorobutadiene in rat liver. Chem. Biol. Interact. 68:1-11.
Werner, M., G. Birner and W. Dekant. 1995a. The role of cytochromeP4503Al/2 in the sex-specific
sulfoxidation of the hexachlorobutadiene metabolite N- Acetyl-
-------
Wolf, C.R., P.N. Berry, J.A. Nash, et al. 1984. Role of microsomal and cytosolic glutathione S-
transferases in the conjugation of hexachloro-1,3-butadiene and its possible relevance to toxicity.
J. Pharmacol. Exp. Ther. 228:202-208.
Woodruff, R.C., J.M. Mason, R. Valencia, et al. 1985. Chemical mutagenesis testing in Drosophila.
V. Results of 53 coded compounds tested for the National Toxicology Program. Environ. Mutagen.
7:677-702.
Yang, R.S. 1988. Hexachloro-1,3-butadiene: toxicology, metabolism, and mechanisms of toxicity.
Rev. Environ. Contam. Toxicol. 101:121-37.
Yang, R.S.H., K.M. Abdo andM.R. Elwell. 1989. Sub chronic toxicology studies of hexachloro-1,3-
butadiene (HCBD) in B6C3FJ mice by dietary incorporation. J. Env. Path. Tox. & One. 9:323-332.
Yip, G. 1976. Survey of hexachloro-1,3,-butadiene in fish, eggs, milk, and vegetables. J. Assoc. Off.
Anal. Chem. 59:559-561.
Yurawecz, M.P., P.A. Dreifuss and L.R. Kamps. 1976. Determination of hexachloro-1,3-butadiene
in spinach, eggs, fish, and milk by electron capture gas-liquid chromatography. J. Assoc. Off. Anal.
Chem. 59:552-558.
Zoeteman, B.C.J., K. Harmsen, J.B.H.J. Linders, et al. 1980. Persistent organic pollutants in river
water and groundwater of the Netherlands. Chemosphere 9:231-249.
HCBD February 2003 10-17
-------
APPENDIX A: Abbreviations and Acronyms
ATSDR
CAS
CCL
CERCLA
CMR
CWS
DWEL
EPA
EPCRA
FDA
GW
HRL
IRIS
MCL
MRL
NAWQA
NCOD
NIOSH
NPDWR
NPL
NTIS
NTNCWS
ppm
PWS
SARA Title III
SDWA
SDWIS
SDWIS FED
SOC
STORE!
SW
TRI
UCM
UCMR
URCIS
U.S. EPA
USGS
VOC
mg/L
>MCL
>MRL
- Agency for Toxic Substances and Disease Registry
- Chemical Abstract Service
- Contaminant Candidate List
- Comprehensive Environmental Response, Compensation & Liability Act
- Chemical Monitoring Reform
- Community Water System
- Drinking Water Equivalent Level
- Environmental Protection Agency
- Emergency Planning and Community Right-to-Know Act
- Food and Drug Administration
- ground water
- Health Reference Level
- Integrated Risk Information System
- Maximum Contaminant Level
- Minimum Reporting Level
- National Water Quality Assessment Program
- National Drinking Water Contaminant Occurrence Database
- National Institute for Occupational Safety and Health
- National Primary Drinking Water Regulation
- National Priorities List
- National Technical Information Service
- Non-Transient Non-Community Water System
- part per million
- Public Water System
- Superfund Amendments and Reauthorization Act
- Safe Drinking Water Act
- Safe Drinking Water Information System
- the Federal Safe Drinking Water Information System
- synthetic organic compound
- Storage and Retrieval System
- surface water
- Toxic Release Inventory
- Unregulated Contaminant Monitoring
- Unregulated Contaminant Monitoring Regulation/Rule
- Unregulated Contaminant Monitoring Information System
- United States Environmental Protection Agency
- United States Geological Survey
- volatile organic compound
- micrograms per liter
- milligrams per liter
- percentage of systems with exceedances
- percentage of systems with detections
HCBD February 2003
A-l
-------
APPENDIX B: Round 1 and Round 2 Occurrence Data Tables for Hexachlorobutadiene
Hexachlorobutadiene Occurrence in Public Water Systems in Round
1, UCM (1987) results
STATE
AK
AL
AR
AZ
CA
CO
DC
DE
FL
GA
HI
IA
IL
IN
KY
LA
MA
MD
Ml
MN
MO
MS
MT
NC
NE
NH
NJ
NM
NV
NY
OH
SD
TN
TX
UT
VI
VT
WA
WV
WY
TOTAL
24
STATE
S
TOTAL
UNIQUE PWS
665
131
448
585
6
1
10
112
127
213
357
524
13
983
1,553
85
297
801
590
8
356
2,655
335
303
2
411
3
992
57
145
12,768
12,284
# GW PWS
540
93
407
571
3
0
8
7
112
149
321
291
9
936
1,529
71
254
790
555
7
252
2,493
306
156
2
391
0
937
26
116
1 1 ,332
10,980
# SW PWS
130
42
47
21
4
1
2
105
16
64
37
233
4
50
28
14
44
11
35
2
123
166
29
147
0
34
3
77
31
38
1,538
1,385
% PWS
with detections
1 .50%
3.05%
0.89%
0.00%
0.00%
0.00%
0.00%
5.36%
0.00%
0.47%
0.00%
0.00%
0.00%
0.10%
0.00%
0.00%
0.00%
0.75%
0.00%
0.00%
0.28%
0.11%
0.30%
0.33%
100.00%
1 .22%
0.00%
0.10%
0.00%
0.00%
0.36%
0.35%
%GW
PWS
with
detections
1 .48%
4.30%
0.74%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.67%
0.00%
0.00%
0.00%
0.11%
0.00%
0.00%
0.00%
0.76%
0.00%
0.00%
0.40%
0.12%
0.33%
0.64%
100.00%
1 .02%
0.00%
0.11%
0.00%
0.00%
0.32%
0.30%
% SW PWS
with
detections
1 .54%
0.00%
2.13%
0.00%
0.00%
0.00%
0.00%
5.71%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
2.94%
0.00%
0.00%
0.00%
0.00%
0.65%
0.72%
% PWS
>HRL
0.00%
1 .53%
0.22%
0.00%
0.00%
0.00%
0.00%
5.36%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.25%
0.00%
0.00%
0.28%
0.08%
0.00%
0.33%
100.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.12%
0.11%
%GW
PWS
> HRL
0.00%
2.15%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.25%
0.00%
0.00%
0.40%
0.08%
0.00%
0.64%
100.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.07%
0.06%
% SW PWS
> HRL
0.00%
0.00%
2.13%
0.00%
0.00%
0.00%
0.00%
5.71 %
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.46%
0.51 %
99% VALUE
(Mkig/L)
< 0.00
0.50
< 2.00
< 6.00
< 0.64
< 0.50
< 0.50
5.00
< 0.30
< 2.00
< 2.00
< 1.00
< 0.50
< 0.50
< 0.50
< 20.00
< 0.50
< 1.20
< 1.00
< 0.20
< 5.00
< 2.00
< 0.50
< 0.50
8.00
< 5.00
< 1.00
< 0.50
< 4.00
< 2.00
< 5.00
< 5.00
PWS = Public Water Systems; GW = Ground Water; SW = Surface Water; MRL = Minimum Reporting Limit (for laboratory analyses);
Health Reference Level = Health Reference Level, an estimated health effect level used for preliminary assessment for this review
The Health Reference Level used for hexachlorobutadiene is 0.9 u(ig/L. This is a draft value for working review only.
Total Number of PWSs = the total number of public water systems with records for hexachlorobutadiene
% PWS with detections, > 1A Health Reference Level, > Health Reference Level = percent of the total number of public water systems with at least one
analytical result that exceeded the MRL, Vi Health Reference Level, Health Reference Level, respectively
99th Percentile Concentration = the concentration value of the 99th percentile of all analytical results (in u(ig/L)
Median Concentration of Detections = the median analytical value of all the detections (analytical results greater than the MRL) (in u(ig/L)
The highlighted states are part of the URCIS (Round 1) 24 State Cross-Section.
HCBD February 2003
B-l
-------
Hexachlorobutadiene Occurrence in Public Water Systems in Round
2, UCM (1993) results
STATE
Tribes (06)
AK
AL
AR
AZ
CA
CO
CT
IN
KY
LA
MA
MD
ME
Ml
MN
MO
MS
NC
ND
NH
NJ
NM
OH
OK
OR
PA
Rl
SC
SD
TN
TX
VT
WA
Wl
TOTAL
20
STATES
TOTAL
UNIQUE PWS
22
625
407
68
14
831
84
117
121
1,310
418
976
744
2,739
1,558
1,412
1
1,775
296
7
720
2,232
790
17
115
237
27
4,412
1
2,548
191
24,815
22,736
# GW PWS
21
481
319
60
11
619
43
107
50
1,241
344
920
676
2,647
1,528
1,297
1
1,585
258
7
693
2,050
541
15
103
216
19
3,825
2,429
188
22,294
20,380
# SW PWS
1
144
88
8
3
212
41
10
71
69
74
56
68
92
30
115
190
38
27
182
249
2
12
21
8
587
1
119
3
2,521
2,356
% PWS
with
detections
0.00%
3.36%
0.00%
0.00%
0.00%
0.24%
0.00%
0.00%
0.00%
0.00%
0.24%
0.20%
0.00%
0.00%
0.00%
0.07%
100.00%
0.51%
0.00%
0.00%
0.14%
0.04%
0.00%
0.00%
0.00%
0.00%
0.00%
0.07%
0.00%
0.00%
0.00%
0.17%
0.18%
% GW
PWS
with
detections
0.00%
2.70%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.11%
0.00%
0.00%
0.00%
0.08%
100.00%
0.44%
0.00%
0.00%
0.14%
0.05%
0.00%
0.00%
0.00%
0.00%
0.00%
0.08%
0.00%
0.00%
0.13%
0.13%
% SW PWS
with
detections
0.00%
5.56%
0.00%
0.00%
0.00%
0.94%
0.00%
0.00%
0.00%
0.00%
1 .35%
1 .79%
0.00%
0.00%
0.00%
0.00%
1 .05%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.56%
0.59%
% PWS
>HRL
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.24%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.04%
0.00%
0.00%
0.00%
0.00%
0.00%
0.05%
0.00%
0.00%
0.00%
0.02%
0.02%
% GW
PWS
> HRL
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.29%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
% SW PWS
>HRL
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.55%
0.00%
0.00%
0.00%
0.00%
0.00%
0.34%
0.00%
0.00%
0.00%
0.12%
0.13%
99% VALUE
(MH9/L)
< 50.00
< 0.00
< 0.10
< 1.00
< 0.50
< 0.00
< 0.00
< 2.00
< 2.50
< 0.50
< 0.50
< 0.50
< 0.00
< 0.00
< 0.50
< 1.00
0.60
< 0.00
< 0.50
< 1.00
< 1.00
< 0.50
< 0.00
< 0.00
< 1.00
< 0.50
< 0.50
1.00
< 0.50
< 0.00
< 0.30
< 1.00
< 1.00
PWS = Public Water Systems; GW = Ground Water; SW = Surface Water; MRL = Minimum Reporting Limit (for laboratory analyses);
Health Reference Level = Health Reference Level, an estimated health effect level used for preliminary assessment for this review
The Health Reference Level used for hexachlorobutadiene is 0.9 (ig/L. This is a draft value for working review only.
Total Number of PWSs = the total number of public water systems with records for hexachlorobutadiene
% PWS with detections, > Vi Health Reference Level, > Health Reference Level = percent of the total number of public water systems with at least one
analytical result that exceeded the MRL, 1A Health Reference Level, Health Reference Level, respectively
99th Percentile Concentration = the concentration value of the 99th percentile of all analytical results (in (ig/L)
Median Concentration of Detections = the median analytical value of all the detections (analytical results greater than the MRL) (in ug/L)
The highlighted States are part of the SDWIS/FED (Round 2) 20 State Cross-Section.
HCBD February 2003
B-2
------- |