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SEPA
United States
Environmental Protection
Agency
Health Effects Support
Document for
Hexachlorobutadiene
External Review Draft
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Health Effects Support Document
for Hexachlorobutadiene
EXTERNAL REVIEW DRAFT
Contract Number: 68-C-01-002
Work Assignment Number: B-02
Prepared for:
U.S. Environmental Protection Agency
Office of Water
Health and Ecological Criteria Division
Washington, DC 20460
Prepared by:
Sciences International, Inc.
1800 Diagonal Road, Suite 500
Alexandria, VA 22314-2808
EPA 822-R-02-028
April 2002
Printed on Recycled Paper
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TABLE OF CONTENTS
LIST OF TABLES v
LIST OF FIGURES ... vi
FOREWORD .... - vii
ACKNOWLEDGMENTS - • - - ix
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 FateinAir ..3-3
3.4 Fate in Water ... 3-3
3.5 FateinSoil 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 4-10
4.3 Conclusions 4-18
5.0 EXPOSURE FROM MEDIA OTHER THAN WATER ........................ 5-1
5.1 ExposurefiomFood 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
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6.1 Absorption 6-1
6.2 Distribution .6-2
6.3 Metabolism 6-2
6.4 .Excretion 6-6
7.0 HAZARD IDENTIFIGATION ... 7-1
• 7.1 HumanEffects .. 7-1
7.1.1 Short-Term Studies 7-1
7.1.2 Long-Tenn and Epidemiological Studies 7-1
7.2 Animal Studies 7-2
7.2.1 Acute Toxicity 7-2
7.2.2 Short-Term Studies 7-5
7.2.3 Subchronic Studies 7-7
7.2.4 Neurotoxicity ..1-9
7.2.5 Developmental/Reproductive Toxicity 7-10
7.2.6 Chronic Toxicity .. 7-11
7.2.7 Carcinogenicity '. 7-12
, 7.3 Other Key Data .... 7-13
7.3.1 Mutagenicity/Genotoxicity • 7-13
7.3.2 Immunotoxicity 7-21
7.3.3 Hormonal Disruption 7-21
7.3.4 Physiological or Mechanistic Studies 7-21
7.3.5 Structure-Activity Relationship „ 7-25
7.4 Hazard Characterization 7-26
7.4.1 Synthesis and Evaluation of Major Noncancer Effects 7-26
7.4.2 Synthesis and Evaluation of Carcinogenic Effects ... 7-28
7.4.3 Mode of Action and Implications in Cancer Assessment 7-32
7.4.4 Weight of Evidence Evaluation for Carcinogenicity 7-33
7.4.5 Sensitive Populations .. 7-34
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 RfCDetermination 8-2
8.2 Dose-Response for Cancer Effects 8-2
8.2.1 Choice of Study '. 8-2
8.2.2 Dose-Response Characterization 8-3
8.2.3 Extrapolation Model and Rationale '..... „. 8-6
8.2.4 Cancer Potency and Unit Risk 8-7
8.2.5 Discussion of Confidence 8-8
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
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9.1.2 National Drinking Water Advisory Council Recommendations ...... 9-2
9.2 Health Effects 9-2
9.2.1 Health Criterion Conclusion i 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 9-7
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 9-10
9.4.3 Relative Source Contribution 9-11
9.4.4 Sensitive Populations 9-12
9.5 Regulatory Determination Decision 9-12
10.0 REFERENCES -. - - - 10-1
APPENDIX A: Abbreviations and Acronyms A-l
APPENDDC B: Round 1 and Round 2 Occurrence
Data Tables for Hexachlorobutadiene B-l
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). ..4-7
Table 4-2. Summary Occurrence Statistics for Hexachlorobutadiene 4-13
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. Mutagenicity of HCBD in Salmonella typhimurium Test Systems 7-15
Table 7-2. Mutagenicity of HCBD Metabolites 7-17
Table 7-3. Genotoxicity of HCBD in Eukaryotic Assay Systems 7-19
Table 7-4. Summary of Principal HCBD Toxicity Studies ... 7-29
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Table 8-1. Incidence of Renal Tubular Neoplasms in Rats Treated with HCBD for
2 Years 8-3
Table 8-2. Dose-Related Changes in the Rodent Kidney after Oral Exposure to HCBD,
Chronic Study - Rat (Kociba et al., 1977) 8-3
Table 8-3. Summary of Cancer Risk Values for HCBD 8-7
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 9-12
Table 9-4. Ratios of Exposures from Various Media to Exposures from
Drinking Water .• 9-12
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)
andRound2 (right) 4-7
Figure 4-2. States with PWSs with Detections of Hexachlorobutadiene for all States
with Data in URCIS (Round 1) and SDWIS/FED (Round 2) 4-15
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 4-16
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) 4-17
Figure 6-1. Proposed Pathways for Hexachlorobutadiene Metabolism 6-3
Figure 8-1. Renal Tumor Dose Response Curves 8-5
External Review Draft—Hexachlorobutadiene—April 2002
VI
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FOREWORD
The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator of
the Environmental Proteetion 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 hi 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 effectsbasis for the preliminary regulatory
determination for hexachlorobutadiene. In arriving at the preliminary 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 wasteful duplication
of effort, information from the following risk assessments by the EPA and other government
agencies were used in development of this document.
U.S. EPA. 1991a. 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-98-00-1
Information from the pubh'shed risk assessments was supplemented with information from
recent studies of hexachlorobutadiene identified by literature searches conducted hi 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
External Review Draft—Hexachlorobutadiene —^ April 2002
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effects such as cellular necrosis. It is expressed, in terms of milligrams per kilogram per day
(mg/kg-day). In general, the RED 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.
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), Guidelines for the
Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986b), Guidelinesfor 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); Useof 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.
External Review Draft—Hexachlorobutadiene—April 2002
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ACKNOWLEDGMENTS
This document was prepared under the U.S. EPA Contract No. 68-C-01 -002. Lead
Scientist, Diana Wong; Ph.D., DABT, Health and Ecological Criteria Division, Office of Science
and Technology, Office of Water.
External Review Draft—Hexachlorpbutadiene — April 2002
IX
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1.0 EXECUTIVE SUMMARY
TKe 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 meanmgfutopportunity for health risk reduction for persons served by public water
systems. EPA will present a determination and further analysis in the Federal Register Notice
covering the Contaminant Candidate List (CCL ) proposals.
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.
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: solubility = 2-2.55 mg/L; vapor
pressure = 0.15 mmHg; Log K^ = 4.78; and Log K^. = 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 fourteen States: AL, AZ, 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 ug/m3 and 0.04 ug/m3, respectively. Air
intake for adults is estimated to be 1.2 x W4 mg/kg-day using the mean air concentration, and is
the main pathway of exposure (U.S. EPA, 1998a). Hexachlorobutadiene is not found in non-fish
dietary foods for the majority of regions of 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) 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. For a health reference level (HRL) of 0.9 ng/L HCBD,
0.114% of the PWSs in Round 1 (74 systems) exceeded the HRL, while 0.018% (11 systems) in
External Review Draft—Hexachlorobutadiene — April 2002
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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.
Hexachlorobutadiene (HCBD) is absorbed following oral administration. Few data are
available on absorption via other routes. 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
which has been implicated in renal toxicity. Evidence exists for a male specific metabolic pathway
in rats (Bimer 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. Studies in
animals show the selective 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. Progressive events over time
include changes in kidney weight, increased excretion of coproporhyrin, renal tubular
degeneration and regeneration, hyperplasia, focal adenomatous proliferation, and renal tumor
formation. Developmental effects were also associated with hexachlorobutadiene exposure in
animals (Harleman and Seinen, 1979). 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). 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 the highest dose of 20 mg/kg-day in the study, benign and malignant tumors were seen
in approximately 23% (9/39) of the male rats, and 15% (6/40) of the female rats. This dose
exceeded the Tna-gitmim 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 a weak carcinogen because it is carcinogenic only at cytotoxic dose..
The nephrotoxicity of HCBD in animals 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. Covalent binding of this reactive thioketene to cellular
macromolecules (e.g. proteins, mitochondrial DNA) and the resultant mitochondria! dysfunction is
believed to contribute to the renal cytotoxicity and tumors observed in animals.
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Limited data from in vitro studies suggest human'have the ability to metabolize HCBD.
However, the activity of HCBD metabolizing enzymes, particularly renal p-lyase, may be many
fold lower in humans than the corresponding enzymes in rats. Thus, there may be less concern for
toxicity in humans.
The primary target for hexachlorobutadiene 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 sensitve to the acute
effects of hexachlorobutadiene (Hook et al., 1983; Lock et al., 1984), suggesting that infants may
also be more susceptible to hexachlorobutadiene toxicity, perhaps as a result of immature organ
systems.
The RfD for hexachlorobutadiene is 2 x HT4 mg/kg-day (EPA, 1998a). It is 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 the NTP (1991) study on mice. A composite
uncertainty factor (UF) of 1,000 was used in the derivation of the RfD. The composite
uncertainty factor included a factor of 10 to account for extrapolation from animals to humans; a
factor of 10 for protection of sensitive subpopulation; a factor of 3 for the use of NOAEL that
may be a minimal LOAEL; and a factor of 3 for database deficiencies (lack of a 2-generation
reproductive study). 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
1996 proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996a), HCBD is
classified as likely to be carcinogenic to humans. Two different approaches were used for dose
response extrapolation to estimate human cancer risk for HCBD from animal data. The linear
approach calculated a slope factor of 4 x 10'2 (mg/kg-day)'1, and a unit risk of 1.1 x 10'5 per ng/L.
Using the nonlinear approach, the point of departure (Pdp) of 0.054 mg/kg-day is the human
equivalent dose for the NOAEL based on absence of renal tubular degeneration and regeneration.
Applying an advisory margin of exposure (MOE) of 300 yields the same value as the RfD. Thus,
the RfD of 2 x 10"4 mg/kg-day will also be protective of cancer effect. In consideration of the
overall evidence, the nonlinear approach was recommended by EPA (1998a).
External Review Draft—Hexachlorobutadiene—April 2002
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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; andGP-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 (U.S. EPA, 1991a). An odor threshold
of 0.006 mg/L has been reported for HCBD in water (U.S. EPA, 1980). HCBD is characterized
by high log K,,,. and log K,,w 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 (f Cl
'Cl
Hexachlorobutadiene
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Table 2-1. Chemical and Physical Properties of Hexachlorobutadiene
Property
Chemical Abstracts Services
(CAS) Registry No.
Chemical Formula
Molecular Weight
Synonyms
NIOSH Registry of Toxic
Effects of Chemical Substances
(RTECS)No.
U.S. EPA Hazardous Waste No.
Oil and Hazardous
Materials/Technical Assistance
Data System (OHM/TADS) No.
Hazardous Substances Data
Bank (HSDB) No.
Boiling Point (at 760 mm Hg)
Melting Point
Vapor Pressure (at 25°Q
Density (at 20°C)
Water Solubility (at 20°C)
Organic Solvents
Partition Coefficients
Odor Threshold (air)
Odor Threshold (water)
Conversion Factor
Information
87-68-3
00.
260.76
HCBD; Perchlorobutadiene;
Hexachlorbutadiene; 1,1,2,3,4,4-Hexachloro-
1 ,3-butadiene; 1 ,3-Hexachlorobutadiene;
Dolen-Pur, GP-40-66:120
EJ0700000
U128
OHM 8100011
2870
215°C
-21°C
0.15 mm Hg
1.55 g/cm3
2-2.55 mg/L
Ethanol, Ether
Log K^ 4.78
LogK^S.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); ChemIDp/i« (2000); HSDB (2000)
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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, and lesser quantities are imported mostly from
Germany as commercial product. HCBD 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 G4 and higher hydrocarbons. The chemical is also used as a fumigant in
Russia, France, Italy, Greece, Spain, and Argentina (ATSDR, 1995; Howard, 1989).
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 jumped 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).
3.2 Release to the Environment
HCBD is listed as a toxic release inventory (TRT) 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 El. 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 is available (1988-1998), surface water discharges generally increased, peaked in
1992-1993, and then decreased significantly through the late 1990s. These 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).
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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
eleven of the Final 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
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industrial needs. HCBD is present in hazardous waste sites in at least 8 States (at NPL sites), has
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 K^ of HCBD indicates that it will readily adsorb to airborne particulate 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.
By analogy to the structurally similar chemical tetrachloroethylene, HCBD is expected to react
with photoehemically-produced hydroxyl radicals and ozone via addition to double bonds
(Atkinson and Carter, 1984; Atkinson, 1987). Mass-balance calculations based on monitoring
data suggest that the half-life of atmospheric HCBD is about 1.6 years in the northern hemisphere
(HSDB, 2000). However, comparison with tetrachloroethylene indicates that the half-life could
be as short as 60 days as a result of reactions with ozone and hydroxyl radicals (HSDB, 2000).
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, 1991a). 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 ng/L, while concentrations in mud
or soil were >200 |ag/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) suggests that this process may occur relatively slowly (U.S. EPA, 1991a).
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).
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The high octanol-water partition coefficient (K^) 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).
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).
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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 the use and environmental release as well as
ambient occurrence of HCBD are also reviewed.
4.1 Ambient Occurrence
Tounderstand 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.
This 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
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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 (Squillace et al., 1999). The minimum reporting limit (MRL) for most VOCs,
including HCBD, was 0.2 ug/L (Squillace et al., 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 (SDWA), 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, 1991c) for required
monitoring that began in 1993 [57 FR 31776] (U.S. EPA, 1992c).
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HCBD has been monitored tinder 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
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 hi a database called the Safe Drinking Water Information System (Federal
version; SDWIS/FED) and can be queried to provide a summary of the data hi SDWIS/FED for a
particular contaminant. The data used hi this report were derived from the data hi SDWIS/FED
and another database called the Unregulated Contaminant Information System (URCIS).
The data hi 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 hi 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 [52 FR 25720]. 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 hi 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 put into a database called the Unregulated Contaminant
Information System, or URCIS. Most of the Phase 1 regulated contaminants were also VOCs.
Both the unregulated and regulated VOCs are analyzed using the same sample and the same
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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
3526]. Monitoring for the UCM (1993) contaminants began coincident with the Phase II/V
regulated 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. While Round 1 includes the period from
1988-1992, it also includes results from samples analyzed prior to 1988 that were
"grandfathered" into the database (for further details see U.S. EPA, 2001a,b).
i •
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 other documents ( U.S. EPA, 2001a,b); readers are referred to these
for more specific information.
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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 iheir 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 HI).
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 IH.A. and IH.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,
2001b, Section DI.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
agrichemical pollution potential rankings and high manufacturing pollution potential rankings;
States with high agrichemical pollution potential rankings and low manufacturing pollution
potential rankings; States with low agrichemical pollution potential rankings and high
manufacturing pollution potential rankings; and States with low agrichemical pollution potential
rankings and low manufacturing pollution potential rankings (U.S. EPA, 200Ib, Section HI.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).
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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
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
HLB.l).
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 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
HI.B.1).
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.
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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).
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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
priority contaminants as described above. These descriptive statistics are summarized here.
B.ased 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 detections) greater than half the Health Reference
Level (URL); or a detections) greater than the Health Reference Level. The Health Reference
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Level, 0.9 ug/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 ug/L—a value twice as large
as the median concentration of detections for Round 1 (0.25 ug/L) (This occurs because some
States and/or systems reporting detections were using a lower MRL and had positive results
lower than the MRL 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
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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 (> V* 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). Between 0.02 and 0.1% of PWSs in Rounds 1 and 2 experienced
detections >HRL, affecting a population of approximately 10,000-780,000.
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%).
External Review Draft—Hexachlorobutadiene—April 2002
4-10
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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 1.12 provided data for HCBD (
U.S. EPA, 2'001a). 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, 0.2-6.4% 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%.
Both Round 1 and Round 2 national cross-sections show a proportionate balance hi
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 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 jig/L. These are not substantively different, given the data sources.
External Review Draft—Hexacklorobutadiene — April 2002
4-11
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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 aliso
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 (TRT)
States for HCBD, nor do they have CERCLA NPL sites with detections of the chemical (Figure
4-4).
External Review Drqfi—Hexachlorobutadiene—April 2002
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Table 4-2. Summary Occurrence Statistics for Hexachlorobutadiene.
, "*•£- \** £v 5 '' t"*' ^ *T| "" ' "
'*(•'..&> " ~, ">"-"' L"
7 ^"£&"? * s •*} *&> ^ i*^ t t
%eque#ey*Faet§fs , TV r ' ^^^ ^ -
Total Number of Samples
Percent of Samples with Detections
99th Percenfile Concentration (all samples)
Health Reference Level
Minimum Reporting Level (MRL)
99°" Percentile Concentration of Detections
Median Concentration of Detections
Total Number of PWSs
Number of GWPWSs
Number of SWPWSs
Total Population
Population of GW PWSs
Population of SW PWSs
"'zrfrStilfe ^
-^o^S^rtan*-
•"-"-(lUnijMi !) '*•
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^tate' ^
'CriBSsSiSeAjon2^
* {Rloiind^
93,585
0.05%
< (Non-detect)
0.9 ng/L
Variable*
,1.5ng/L
0.30 \ig/L
22,736
20,380
2,356
67,075,493
24,960^22
42,115,271
Occurrence by System
% PWSs with detections (> MRL)
Range of Cross-Section States
GW PWSs with detections
SW PWSs with detections
% PWSs > 1/2 Health Reference Level (HRL)
Range of Cross-Section States
GW PWSs > 1/2 Health Reference Level
SW PWSs > 1/2 Health Reference Level
% PWSs > Health Reference Level
Range of Cross-Section States
GW PWSs > Health Reference Level
SW PWSs > Health Reference Level
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%
0.180%
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%
1* ^ ,-
.'< "* *',-«•'*'• ^
National Sy stem & j
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 2
117
N/A
79
33
51
N/A
38
12
11
N/A
3
7
External Review Draft—Hexachlorobutadiene—April 2002
4-13
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Table 4-2 (continued)
Frequency Factors
Occurrence by Population Served
% PWS Population Served with detections
Range of Cross-Section States
GW PWS Population with detections
SW PWS Population with detections
% PWS Population Served > 1/2 Health Ref Level
Range of Cross-Section States
GW PWS Population > 1/2 Health Ref
Level
SW PWS Population > 1/2 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!)
0.896%
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-Stite; '
^Cross-Section*
(Round 1}
2.360%
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%
' -'" *''< ''; t*f
f National System &
Population Numbers*
Round 1
1,909,000
N/A
1,249,000
194,000
1,213,000
N/A
838,000
46,000
• 781,000
N/A
531,000
46,000
Round 2
5,027,000
N/A
159,000
4,646,000
4,965,000
N/A
152,000
4,5S>3,000
10,000
N/A
9,000
1,000
1. SummnryResultsba5edondatafrom24-StateCross-Section,fromURCIS,UCM(1987)Round 1.
Z Summary Results based on data from 20-State Cross-Section, from SDWIS/FED, UCM (1993) Round 2.
3. Total PWSnnd population numbers are fixim 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 ug/L. This is a draft value for working review only.
- Total Number of Samples = the total number of analytical records for hexachlorobutadiene.
- 99th Pcrcentile Concentration = the concentration value of the 99th percentile of either all analytical results or just the samples with detections (in
ug/L).
- 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 > Vi 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, 'A Health Reference Level, Health Reference Level, respectively.
- % PWS Population Served with detections, % PWS Population Served >!/i 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, Vi Health Reference Level, or the Health
Reference Level, respectively. . -
External Review Draft—Hexachlorobutadiene — April 2002
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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
I States not in Round 1 or Round 2
55H No data for Hexachlorobutadiene
Sll States with No Detections (No P WSs > MRL)
JH| States with Detections (Any PWSs > MRL)
External Review Draft—Hexachlorobutadiene — April 2002
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Figure 4-3. States with PWSs with Detections of Hexachlorobutadieme (any PWSs with
results greater than the Minimum Reporting Level [MRL]) for Round 1
(above) and Round 2 (below) Cross-section States.
• SMoofFtoriiUU « outlier will 5.36% PWS > MRL
HexacMorobutadlene Occurrence IB Round 1
I I States not in Cross-Section
r"™J No data for Hcxachloibatadicnc
EMI 0.00% PWSs > MRL
0.01 - 1.00% PWSs > MRL
mtm 1.00-3.50% PWSs > MRL'
HexacBlorobutadieBe Occurrence
IB Round 2
I States nol In Crass-Section
~""~\ No data for Hexachloibutadiene
gH 0.00'/. PWSs > MRL
• 0.01 -1.00% PWSs > MRL
SI 1.00 -3.50% PWSs > MRL
External Review Draft —HexaMorobutadiene—April 2002
4-16
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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).
* Stt« of north a an outlfer with 5.36% PWS > MRL
Hexacblorobiitadlene Occurrence
in Roynd 1 »d Round 2
1 States not in Cross-Section
j No data for Hexachforbutadiene
m 0.00% PWSs > MRL
I 0.01 -1.00% PWSs > MRL
I 1.00-350%PWSs> MRL*
• Saw omoricU u an outlier with 5.36% PWS > HRL
HencbiorobotadieBe OccurrcBce
In Ronid 1 and Round 2
States not in Cross-Section
No data for Hexachtorobutadiene
0.00% PWSs > HRL
0.01-1.00% PWSs > HRL
1.00 -3.50% PWSs > HRL*
External Review Draft—Hexachlorobutadiene—April 2002
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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 hi 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 p,g/L and the 99th percentile
concentration is 10 ng/L. Median and 99th percentile concentrations for Round 2 detections are
0.30 ug/L and 1.5 ug/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).
External Review Draft—Hexachlorobutadiene—April 2002
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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 artifiact.
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 in
Louisiana were 226.33 ± 778.40 ng/g hi fish collected from a contaminated study site 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;
De Vault, 1985), with the exception of trout from Lake Ontario, which were reported to contain
60 to 300 ng/g (Oliver and Nimi, 1983). HCBD was not detected in 51 biota samples catalogued
in the STORET database (Staples et al., 1985).
External Review Draft—Hexachlorobutadiene—April 2002
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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.1 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 (Dreissena polymorphd)
and eel (Anguilla anguittd) 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.
1 The total number of samples for this study is not clear. If one sample of bottom-feeding fish was taken from all
sites, then the total number of composite samples was most likely 468.
External Review Draft—Hexachlorobutadiene—April 2002
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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)
5.13 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, although HCBD may have been incorrectly measured in milk 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 feet 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 * 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 majority 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.
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
External Review Draft—Hexachlorobutadiene—April 2002
5-3
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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 Ruehl et al.
(1994) data by a fish intake of 18 g/day for the general population and dividing by a body v/eight
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 422 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 tune 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
concentrations measured by Shah and Heyerdahl (1988) were 0.036 parts per billion (ppb) (0.42
jig/m3) and 0.003 ppb (0.04 u.g/m3), respectively. The 25th and 75th percentiles were 0.001 ppb
(0.01 Hg/m3) and 0.006 ppb (0.07 u.g/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
[ig/m3). Of 16 samples taken from source-dominated areas, the median was 0.002 ppb (0.02
u.g/m3). No indoor concentrations were reported (Shah and Heyerdahl, 1988).
Shah and HeyerdahFs compilation included a study conducted by 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
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 positive, with concentrations of 18 and 37 jig/m3. Sampling in Houston, Texas, and
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surrounding areas showed a range of trace levels to 2,066 p-g/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 p.g/m3) HCBD at 18
locations sampled from 1982 to 1986. The detection limits in mis 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
jig/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 jig/m3.
A number of cities had HCBD levels ranging from 2 to 11 ppt (0.02 to 0.12 jig/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
jig/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 fairly large number
of observations (n=72). For adults, the mean concentration of 0.42 |J.g/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 |j.g 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 m3/day (U.S. EPA, 1988). The resulting value was divided by a body
weight of 10 kg and the units were converted from |j.g to mg, resulting in an intake of 6.3 x 1Q"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 is not indicated.
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5.3 Exposure from Soil
5.3.1 Concentration of HCBD in Soil and Sediment
No data were located on the concentration of HCBD in either soil or dust However,
sediments have been shown to 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 jig/kg HCBD (Oliver and Bourbonniere,
1985). Sediments from the Great Lakes were reported to contain levels of HCBD typically
ranging from 0.08 to 120 jig/kg (McConnell et al., 1975). Recent data for suspended solids
collected from the Rhine-Meuse river basin indicate HCBD levels ranging from < 3.4 to 19
(Hendriks et al., 1998). This concentration range is comparable to the earlier data reported 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 ng/kg to 0.40 ng/kg (Abdelghani et al,. 1995). These concentrations were well below the
action levels of 4,000 jig/kg for sediment (U.S. EPA, 1991a). At a Federal Superfund site near
Baton Rouge, Louisiana, preliminary data from a sampling of sediments showed HCBD levels
from 2 to 3,770 nag/kg (U.S. EPA, 1992b). The HCBD level in a sediment sample from Lake
Charles, Louisiana was found to be 3,500 ng/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 ng/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 jig/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 |J.g/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 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 to HCBD from non-fish dietary sources, inspection of the data indicates that most
intake of HCBD by the general population occurs via respiratory intake. However, it should be
cautioned that this determination is subject to a number of uncertainties: 1) the database for the
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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): 0.6 ng/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'xlQ-4
Child
6.3 x 10"4
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 hi soil and dust were not available to estimate
via this pathway.
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6.0 TOXICOKINETieS
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 m humans.
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. Reichert et.al. (1985) administered 1 mg/kg of 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). 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
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.
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6.2 Distribution
HCBD has been detected in human blood (Bristol et al.,. 1982) and in the adipose tissue of
human cadavers (Mes et al., 1985). 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 |*g 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 and its metabolites preferentially distributed to the
kidney, liver, adipose tissue and brain of experimental animals (Reichert, 1983; Reichert et al.,
1985; Dekant et al., 1988a). Covalent binding of HCBD-related radioactivity to tissue proteins
was highest during the first six hours after dosing, and was higher in the kidney than in the liver;
and this effect was independent of dose (Reichert et al., 1985). In rats administered 1 mg/kg 14C-
HCBD, covalent binding of the radioactivity to protein in the kidney was about twice mat in the
liver 72 hours after dosing (Reichert et al., 1985). Nash et al. (1984) reported a specific
localization of administered radioactivity in 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 rats in which the bile ducts of one group of animals administered an oral
dose of 100 mg/kg "C-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 hi humans or annuals after
inhalation or dermal exposure. Davis et al. (1980) administered 0.1 mg/kg radiolabeled 14C-
HCBD as a tracer dose to male Sprague-Dawley rats (5 animals/group) via intraperitdneal
injection. A nephrotoxic group received the same amount of labeled HCBD plus 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 admmisteired 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 hi 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 (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, HCBD is initially 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). In rats, a di-substituted glutathione conjugate, 1,4 bis(ls2,3,4-
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 and
Dekant, 1992). Nash et al. (1984), for example, collected bile excretions from cannulated rats
that had been orally ,
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administered 200 nag/kg 14C-HCBD and determined that 40% of the bile radioactivity was
associated with the glutathione conjugate. The weight of evidence suggests that oxidative
reactions involving cytochrome P450 have little role in the initial metabolism of HCBD (Wolf et
al., 1984; Dekant etal.,1988a).
The glutathione conjugate of HCBD can be reabsorbed intact from the gastrointestinal
tract (Koob and Dekant, 1992; Gietl et al., 1991). Alternatively, a portion of it can b? catabolized
by Y-glutamyltranspeptidase and dipeptidases in the gastrointestinal tract to the cysteine
conjugate, ^-(l,l,2,3,4-pentachlorobutadienyl)-L-cysteine (Jones etal., 1985; Gietl et al., 1991;
Koob and Dekant, 1992). 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 et
al., 1991; Gietl and Anders, 1991), or be excreted with the feces (Dekant et al. 1988a). However,
the majority of the glutathione conjugate is delivered to the kidney by systemic circulation (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 liver, a minor fraction of this metabolite is converted
by 7V-aceryltransferase to an acetylated cysteine conjugate, N-acety\-S-(l, 1,2,3,4-
pentachlorobutadienyl)-L-cysteme (AT-AcPCBC) (Koob and Dekant, 1992).
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 electropbilic
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, N-acetyl-S-(l,l,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-
pentachlorobutadiene methylthioether and 1,1,2,3,4-pentachlorobutadiene carboxymethylthioether
(Reichert et al., 1985). In addition, 1 to 8% of the administered radioactivity is oxidized to
carbon dioxide in rats (Reichert et al., 1985; Payan et al., 1991).
Evidence for a male-specific HCBD metabolic pathway in rats has been reported by Birner
and colleagues (Birner et al., 1995,1998; Werner et al.!995a). The metabolite N-acetyl-S-
(l,l,2,3,4-pentachlorobutadienyl)-L-cysteine sulfoxide (AT-AcPCBC-SO) is detected in the urine
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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 (Bimer et al., 1995; Werner et al., 1995a). This metabolite has been found to be
cytotoxic to proximal tubular cells in vitro without activation by jJ-lyase (Birner et al., 1995).
When given intravenously, N-AcPCBC-SO produced necroses of the kidney tubules in male rats
(Birner et al., 1998).
The N-AcPCBC-SO formed in male rats occurs as two diastereomers present in equimolar
amounts: (K)-N-AcPCBC-SO and (5)-JV-AcPCBC-SO (Werner et al., 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 S-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 (/?)-^T-
acetyl-Sr-(4-glutathion-iS'-yl-l,2,3,4-tetrachlorobutadienyl)-L-cysteine sulfoxide. Incubation of the
(5)-sulfoxide produced two glutathione conjugates identified as (S)-N-acetyl-,S'-(4-glutamion-,S'-yl-
l,2,3,4-tetrachlorobutadienyl)-L-cysteine sulfoxide and (5)-7V-acetyl-5'-(2-glutathion-5-yl-l,3,4,4-
tetrachlorobutadienyl)-L-cysteine sulfoxide. In the presence of rat kidney cytosol, only the (5)-J\T-
acetyl-5-(2-glutathion-iS'-yl-l,3,4,4-tetrachlorobutadienyl)-L-cysteine sulfoxide conjugate was
formed. Glutathione conjugation of the (jR)-sulfoxide was not observed. The observed pattern
of product formation was attributed to catalysis by different glutathione S-transferases in liver (a-
and n-class) and kidney (a-class). This hypothesis was confirmed by product analysis following
incubation of N-AcPCBC-SO with purified rat a- and ^.-class glutathione iS'-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)glutatbione (PCBG) from HCBD in
human liver cytosol is approximately 20 to 30% of the rates observed in rat and mouse cytosol
(Dekant et al. 1998).
The enzyme Y-ghitamyl 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).
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Werner et al. (1995b) demonstrated that human liver microsomes are capable of oxidizing
^jS^pentachlorobutadieny^-L-cysteirie to the corresponding sulfoxide (N-
AcPCBC-SO). In contrast to the male-specific formation of AT-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 3 A 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 N-AcPCBC-SO with purified human glutathione S-
transferase Ml-1 (|i-class) catalyzes the formation of (
-------
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 adtoinistered 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 ratsor mice include S-(l, 1,2,3,4-
pentachlorobutadienyl) glutathione, S-(l,l,2,3,4-pen1achlorobutadienyl)-L-cysteine, N-acety\-S-
(l,l,2,3,4-pentachlorobutadienyl)-L-cysteine, 1,1,2,3,4-pentachlorobutadienyl sulfenic acid,
1,1 j2,3,4-pentachlorobutadiene methylthioether, 1,1,2,3,4-pentachlorobutadiene
carboxymethylthioether, and 1,1,2,3-tetrachlorobutenoic acid (Dekant et al., 1988a; Nash et al.,
1984; Reichert and Schultz, 1986; Reichert et al., 1985). As noted previously, the novel
metabolite A/-acetyl-5-(l,l,2,3,4,-pentachlorobutadienyl)-L-cysteine sulfoxide has been detected in
the urine of male, but not female, rats following oral administration of HCBD (Bimer et al.,
1995)..
Comparatively few data are available on the identity of fecal metabolites. Dekant etal.
(1988a) administered a single 30 mg/kg gavage dose of 14C-HCBD in com 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 iS-(l,l,2,3,4-pentachlorobutadienyl)
glutathione. The remainder of the fecal radioactivity was present as polar metabolites which
could not be structurally identified.
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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, the 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 smell functions (U.S. EPA,
1991a). Interpretation of these data is confounded by concurrent exposure of the 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).
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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. Single-dose oral LD50 values for adult rats ranged from 200 to
400 mg/kg for females, and 504 to 667 mg/kg for males (unpublished study cited in Schwetz et
al., 1977). 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, 1991a).
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 increased levels of glucose,
protein, alkaline phosphatase, Af-acetyl^-D-glucosaminidase (NAG), y-glutamyl transpeptiidase
and alanine aminopeptidase. These biochemical changes were indicative 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
com 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 "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 24 hours after exposure, and the right kidneys were subjected to microscopic
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examination. Nephrotoxicity was also evaluated by determination of urinary p2-microglobulin, as
well as 7-glutamyl transpeptidase, aspartate aminotransferase (AST), and NAG activity. A
Lowest-Observed-Adverse-Effect Level (LOAEL) of 100 nig/kg was identified in this study on
the basis of kidney lesions and a three-fold higher urinary 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 et al. (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 OF1 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).
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
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animals from the two highest d
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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 N-
acetylcysteine 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.
7.2.2 Short-Term Studies
Oral exposure
Kociba et al. (1971) conducted an unpublished Dow Chemical Company study of HCBD
toxicity, the results of which were summarized in 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-
Observed-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
com 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 National Toxicology Program (NTP) conducted a 2-week oral exposure study in
B6C3F, mice (NTP, 1991; Yang etal., 1989). Groups of mice (5 animals/sex/dose group)
received diets containing nominal concentrations of 0, 30,100, 300,1,000 or 3,000 ppm HCBD
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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 300 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 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 lethal dose groups.
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 lethal 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
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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. 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; 1,067 mg/m3 for 12 days, 6 hours/day; or 2,668 mg/m3 for 2 days, 4 hours/day.
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. 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 and female 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 or 20 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. These
results indicate a NOAEL of 0.2 mg/kg-day for male and female rats, based on the absence of
renal histopathology or other toxic effects at this dose.
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Harleman and Seiuen (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 cytoplasmic basopbilia 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.
NIP (1991) conducted a 13-week oral exposure study hi B6C3F, 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) hi 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 hi this organ.
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 basopbilic 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. Lesions
were observed hi male mice at dose levels of 4.9 mg/kg-day and above (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 a NOAEL 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 a NOAEL 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).
Nakagawa et al. (1998) administered 0.1% HCBD hi the diet to male Wistar rats (21 rats/
group) for 30 weeks hi 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
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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 B6C3F, 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 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.
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 (WHO, 1994).
Increased levels of free radicals were detected in the brain and spinal cord of the offspring of
treated dams (U.S. EPA, 1991a).
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.
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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 were observed on pregnancy rate, time to delivery, neonatal survival, neonate sex ratio,
weanling histopathology, or incidence of neonatal external, visceral, or skeletal anomalies.
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, no
conceptions occurred, the ovaries showed little follicular activity, and no uterine implantation sites
were observed at the 150 mg/kg-day dose. 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 days 0,10, and 20. No gross abnormalities were noted. The LOAEL
identified was 15 mg/kg-day. i
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, 1991a). Histological changes in the nerve cells and myelin
fibers of the spinal cord were noted in both dams and offspring (WHO, 1994). 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, 1991a).
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 thejaverage
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).
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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. 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.
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.
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 (Hardin et al., 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, 1991a).
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, and serum clinical chemistry, and histopathology of major organs. The investigators
reported significantly increased mortality in high-dose males (p<0.05). 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).
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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 inpreased
urinary coproporphyrin excretion (females only), 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.
Microscopic examination revealed histopathological changes in the kidney, including hyperplasia
and neoplasia of the renal epithelium. 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 arid 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 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. 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.
Chudin et al. (1985) conducted a 1-year HCBD oral exposure study in rats. The average
daily doses ranged from 0.6 to 37 mg/kg-day. Benign liver and kidney tumors were noted, but
malignant tumors were not observed (U.S. EPA, 1991a).
Nakagawa et al. (1998) investigated the effect of HCBD on renal carcinogenesis in male
Wistar rats pre-treated with JV-ethyl-JV-hydroxyethyhiitrosamine (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 of EHEN. HCBD was administered for 30 weeks at a concentration of
0.1% by weight (1,000 mg/kg) in the diet to rats (12/treatment group) that had previously
received 0.1% EHEN in the drinking water. The combined treatment with HCBD and BEEN
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.
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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.
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 hi 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 hi 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, and the limited scope of
histopathological evaluation (WHO, 1994).
7.3 Other Key Data
7.3.1 Mutagenicity/Genotoxicity
The mutagenicity of HCBD has been evaluated hi 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 hi humans.
Bacterial Test Systems ... • T
Test results from bacterial assays of mutagenicity are summarized in Table 7-1. Most tests
in standard 5. 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 hi Salmonella typhimurium was reported in the presence of metabolic activation
induced by rat liver S9 fraction (U.S. EPA, 1991a). Conflicting results hi standard assays may be
due to contaminants hi technical and even analytical grade HCBD (Reichert et al., 1984;
External Review Draft—Hexachlorobutadiene — April 2002
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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%5 a negative mutagenic
response was obtained.
Positive results have been reported for HCBD when specialized pre-incubation conditions
that included rat liver microsomes and glutatbione were utilized (Vamvakas et al., 1988; Roldan-
Arjona et al., 1991). The mutagenic response was increased with additional inclusion of rat
kidney microsomes as a y-glutamyl transpeptidase and dipeptidase source (Vamvakas et al.,
1988). Reichert et al. (1984) also 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 7-2.
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, JV-acetyl-S-pentachlorobutadienyl-L-cysteine (Reichert and
Schutz, 1986; Wild et al., 1986). Other 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). 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
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Table 7-2. Mutagenicity of HCBD Metabolites
Metabolite8
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
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+ ratliver S9
-f rat liver S9
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+ rat liver S9
+ rat liver S9
+ rat liver S9
Result*
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Reference
Green & Odum (1985)
Green & Odum (1985),
Vamvakas et al. (1988)
Vamvakas et al. (1 988)
Vamvakas et al. (1988)
Green & Odum (1985)
Dekant et al. (1986)
Wild etal. (1986)
Reichert and Schutz
(1986)
Wild etal. (1986)
Wild etal. (1986)
Reichert etal. (1984)
Reichert et al. (1984)
Source: modified from WHO (1994)
"Abbreviations: PCBG S-(l,l,2,3,4-pentachlorobutadienyl) glutathione
TCBG 1,4 bis(l^,3,4-tetrachlorobutadienyl) glutathione
TCBC l,4bis(l,2,3,4-tetrachlorobutadienyl>L-cysteine
PCBC 5-(l,l,2,3,4-pentachlorobutadienyl)-L-cysteine
Af-AcPCBC Atacetyl-5'-(l,l,2,3)4-pentachlorobutadienyl)-L-cysteine
PCCMT. 1,1,2,3,4-pentachlorobutadiene carboxymethylthioether
PCMT l,l^,3,4.pentachlorobutadienemethylthioether
PCBA , 2,2,3,4,4-pentachloro-3-butenoic acid
PCBAC 2^,3,4,4-pentachloro-3-butenoic acid chloride
b S9* = a fortified S9 mix containing 3 times the normalprotein concentration; GSH = reduced glutathione
c+ = 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 p-lyase inhibitor aminooxyacetic acid
External Review Draft—Hexachlorobutadiene—April 2002
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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 mammaliac test systems are summarized in the upper
portion of Table 7-3. 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 £-(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 mat renal metabolism plays a key role in PCBG-induced
genotoxicity.
In Vivo Test Systems
Results of in vivo HCBD genotoxicity tests are summarized hi the lower portion of Table
7-3. Both negative (NIOSH, 1981; Schwetz et al., 1977) and positiye (German, 1988) results
have been reported for chromosome aberration assays conducted in HCBD-treated mice or rats.
Negative findings have been reported hi a dominant lethal assay hi rats (NIOSH, 1981), and hi the
Drosophila melanogaster sex-linked recessive lethal mutation assay with exposure via either
injection 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) hi kidney cells from rats fed 20 mg/kg-day HCBD hi the diet for 3 weeks,
suggesting mat HCBD exhibited a minor degree of renal genotoxicity.
Schrenk and Dekant (1989) evaluated the covalent binding of HCBD metabolites to renal
and hepatic DNA hi NMRI mice. A low level of covalent binding (covalent binding index (CBI) =
27) was observed hi nuclear DNA (nDNA) isolated from the kidney, while covalent binding was
undetectable hi nDNA isolated from liver. Significantly higher levels of covalent binding were
observed hi 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.
External Review Draft—Hexachlorobutadiene—April 2002
7-18
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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
was observed in B6C3F, mice administered lethal doses 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). Performance of immune function
screening batteries in HCBD-treated animals has not been evaluated,.,
Delayed hypersensitivity reaction was exhibited hi guinea pigs to dermal HCBD
application (Gradiskiet al., 1975).
7.3.3 Hormonal Disruption
No studies were identified that associate HCBD exposure with endocrine disruption.
7.3.4 Physiological or Mechanistic Studies
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 (Dekant et al., 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 unmunocytochemical technique that the region of highest cytosolic p-lyase
activity in untreated rats coincides with the site of HCBD-induced necrosis hi 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 S,-S2
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 hi the renal cortex shortly after dosing. Renal
External Review Draft—Hexachlorobutadiene—April 2002
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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 glutathione conjugate to 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 16 or 64 umol/kg 14C-radiolabeled JV-acetyl-pentachlorobutadienyl-L-
cysteine to female Alpk/AP rats and observed acute renal necrosis within four hours. Prior
administration of up to 500 umol/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 the protective effect of probenecid against HCBD
metabolite-induced toxicity in freshly isolated 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 N-acetyl-pentachlorobutadienyl-L-cysteine to the incubation
medium reduced 3H-proline incorporation to 34% of the control value. The inclusion of 400 uM
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 mitochondria! compartment, an elevation of cytosolic Ca2+
concentration, inhibition of respiration, and decreased levels of ATP. Prevention of PCBG
bioactivation by inhibition of Y-ghitamyl 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.
WalHn et al. (1987) studied S-pentachlorobutadienyl-L-cysteine (PCBC) toxicity in
mitochondria isolated from the rat kidney cortex. Respiring mitochondria exposed to PCBC
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showed a dose-dependent loss of ability to retain calcium. This effect was associated with a
collapse of mitoehondrial membrane potential. A slow nonenzymatic depletion of mitochondrial
glutathione was also observed. Preincubation with amindoxyacetic 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 PCBG-induced toxicity in renal
proximal tubules isolated from New Zealand rabbits. Addition of 20 to 500 \iM PCBC to renal
tubule suspension initiated a specific sequence of toxic effects. Fifteen minutes of exposure to .
200 uM PCBC caused an increase in basal and ouabain-insensitive respiration. Sixty minutes of
PCBC exposure inhibited basal, nystatin-stimulated and ouabain insensitive respiration, and
resulted in a 79% decrease in glutathione concentration. In addition, an 11% decrease in lactate
dehydrogenase retention was observed after 60 minutes of PCBC exposure, suggesting that cell
viability was decreased as a result of treatment. Analysis of mitochondrial function indicated that
the initial increase in respiration resulted from uncoupling of oxidative phosphorylation. The
resulting ATP deficiency may have limited energy-dependent active transport processes hi the
tubules, thus inhibiting reabsorption processes. The changes in respiration observed at 60 minutes .
appeared to result from gross mitochondrial damage characterized by inhibition of state 3
respiration, depression of cytochfome 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 et al. (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 etal. (1991) demonstrated the rapid
accumulation of 35S-PCBC hi renal proximal tubule cells and its metabolism to a reactive
intermediate mat bound to tubular protein. An estimated 70 to 90% of the mtracellular
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 subc'ellular 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 lipoyf 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.
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Schrenk and Dekant (1989) investigated covalent binding of 14C-labeled HCBD
metabolites to mouse DNA after a single oral dose of 30 mg/kg 14C-HCBD. HCBD metabolites
boimd 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 tor cysteine
conjugate p-lyase. The activity of p-lyase leads to formation of an enethiol intermediate wMch 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-
methylthioeth'er. Morel et al. (1999) investigated the role of S-adenosyl methionine (SAM)-
dependent thiol methylation in prevention of HCBD-induced nephrotoxicity in male Swiss OF1
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 mg HCBD/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. 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 a2ll-globulin nephropathy represents a potential alternative mechanism
for HCBD toxicity in rats. Since oc^-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 o^-globulin is a prerequisite
for development of nephrotoxicity. Because HCBD-induced renal toxicity occurs in'both male
and female rats, it is evident that o^-globulin nephropathy is not required for nephrotoxicity.
However, there is limited evidence to suggest that the a2ll-globulin mechanism may contribute to
the nephrotoxicity observed in male animals. Bimer et al. (1995) observed that unmetabolized
HCBD was excreted hi 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 a2ll-globulin-deficient strain. 14C4HCBD 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 «2(l-
globulin. Histopathological examination 48 hours after treatment revealed the formulation of
hyaline droplets indicative of (^-globulin accumulation in renal epithelial cells. In addition,
microscopic examination confirmed the occurrence of more extensive nephropathy in male: than in
female animals as previously observed in Wistar rats.
Saito et al. (1996) established that dose-dependent levels of kidney-type a2(1-globul:in (ccG-
K) in the urine are a reliable predictor of a2(1-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 ccG-K was detected following exposure,
suggesting that HCBD treatment did not induce a marked accumulation of o^-globulin.
However, histopathological examination revealed some epithelial cells showing hyaline droplet-
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related degeneration. The size of the hyaline droplets formed following HCBD-treatment were
generally smaller than those observed after treatment with the well-characterized a^-globulin
nephropathy-inducing agent cWimonene. =e..
No significant increase in a2(1-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 (PShler et al., 1999).
7.3.5 Structure-Activity Relationship
HCBD appears to share a common mechanism of toxicity with several structurally-related
haloalkenes, including perfluoropropene, trichloroethene, tetrachloroethene, and
trichlorotrifluoropropene. All of these chemicals are selectively toxic to the proximal tubule. The
common basis for toxicity is bioactivation of these compounds by a multistep pathway which is
initiated by conjugation with glutathione, resulting in the formation of a glutathione S-conjugate.
Metabolism to the corresponding cysteine S-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 mitpchondrial macromolecules,
resulling 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 it-bond.
Chloroalkenes are thus reported to be more resistant to metabolism by glutathione conjugation
than are fluoroalkenes.
Multiple investigations have compared the toxicity of structurally-related haloalkene
conjugates. Anders et al. (1987) and Lock (1988) evaluated the toxicity of structural analogues
of HCBD conjugates and noted remarkable similarity. 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
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activated by p-lyase and had a similar effect on ion transport, their mutagenicity differed. The
conjugates of HCBD, trichloroethylene andperchloroethylene were mutagenic in the Ames
bacterial mutation assay when activated by rat kidney S9 fraction. In contrast, the conjugates of
TFE and HEP were not mutagenic either in the presence or absence of rat kidney S9 fraction.
Birner et al. (1997) compared the nephrotoxicity of cysteine S-conjugates derived from
trichloroethene, tetrachloroethene, and HCBD. Male and female rats received identical
intravenous doses of 5-(l,2-dichlorovinyl)-L-cysteuie (1,2-DCVC), 5-(2,2-dichlorovinyl)-L-
cysteine (2,2-DCVC), S-(l,2,2-1richlorovinyl)-L-cysteine (TCVC), or £-(1,2,3,4,4-
pentachlorobutadienyl)-L-cysteine (PCBC). Assessment of the relative nephrotoxic potency of
the conjugates by histopathological examination and excretion of y-glutamyltranspepidases hi
urine indicated a decrease hi 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 reliable reports of human health effects following HCBD exposure by any
route. Oral exposure studies of HCBD toxicity in animals are summarized hi Table 7-4. A
distinctive feature of HCBD toxicity hi 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 over time include changes in kidney weight, increased excretion of
coproporphyrin, renal tubular degeneration, necrosis and regeneration, hyperplasia, focal
adenomatous proliferation, and tumor formation. Tumor formation occurs exclusively hi 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-S-transferase mediated biosynthesis of a glutathione conjugate
(PCBG) hi the liver. After elimination into the bile, PCBG undergoes subsequent metabolism to a
cysteine conjugate (PCBC) hi the bile, gut or kidneys. PCBC may be acetylated by renal N-
acetyltransferases to form aN-acetyl cysteine conjugates (JV-AcPCBC). Both PCBC and N-
AcPCBC are concentrated hi renal cells via an active transport system (Dekant, 1990). N-
AcPCBC can be excreted hi the urine or de-acetylated to regenerate PCBC. PCBC is a substrate
.for p-lyase-dependent activation to a highly reactive thioketene hi the kidney. Covalent binding of
this reactive species to cellular macromolecules is believed to initiate the damage that ultimately
results ha 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 hi 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 hi this segment of the proximal
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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 mitochondria! membranes. The
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.
. . . - - ' : ij- . -\.l
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
HGBD in male rats. Current evidence suggests that at least two discrete male-specific pathways
may participate hi the more pronounced necrotic changes observed in the renal tubules of male
rats in some studies. Formation of hyaline droplets indicative of a^-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 3A-mediated formation of an JV-acetylated cysteine
conjugate sulfoxide. .• n
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 another study, 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. 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 action for these effects have 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 is
probably related to the distribution of HCBD and its reactive 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
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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. However, the rate of metabolism, particularly for the reaction
catalyzed by p-lyase, appears to be much lower than that observed in rats.
7.4.2 Synthesis and Evaluation of Carcinogenic Effects
No studies of 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 were also observed. 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 hi Ihe outer stripe and cortex of kidneys from HCBD-treated rats. '•.
However, these data must be considered as too limited to support a conclusion of 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 fihe
database include the absence of cell proliferation studies and limited in vivo data for mutagenesis.
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 HCBD
metabolites can bind to DNA in vivo in mice (Schrenk and Dekant, 1989) strengthens this
conclusion.
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7.43 Mode of Action and Implications in Cancer Assessment
i
Both sustained cytotoxic damage and irreversible DNA binding have been proposed as
events in HCBD carcinogenesis (Stott et al., 1981). An oral carcinogenic 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; Reichert et al., 1984; Reichert and Schutz,
1986; Wild et al., 1986), but not all, studies (See Section 7.3.1). Thus, a mutagenic mode of
action cannot be ruled out (Dekant et al., 1990; Lock, 1994).
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. Interaction of highly reactive metabolites with components of the
mitochondrial inner membrane, such as enzymes related to cell function, may result in
mitochondrial dysfunction. 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
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
important proteins involved in the oxidative phosphorylation, 22 transfer RNAs (tRNA) and 2
ribosomal RNAs (rRNA), functional mtDNA is critical to the normal function of a pell.
External Review Draft—Hexachlorobutadiene—April 2002 \ 7-32
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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
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 for HCBD-mediated carcinogenesis. Thus, HCBD induced cytotoxicity and
tumorigensis may be ultimately the consequence of mitochondrial dysfunction resulting from
exposure. . '
Recent evidence suggests that HCBD-induced o^-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, o^-
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 hi 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, 1996a), HCBD is classified as likely to be carcinogenic to humans.
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.
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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 mat
infants may represent a potentially sensitive subpopulation for acute HCBD exposure, perhaps as
a result of immature organ systems.
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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 NOAEL/LOAEL approach.
Choice of Principal Study and Critical Effect
There are no reliable dose-response data for humans exposed to HGBD. The Kociba et al.
(1977) study on rats and the NTP (1991) study on mice are chosen as co-principal studies. The
RfD for HCBD is derived from a NOAEL of 0.2 mg/kg-day for renal tubular epithelial cell
hyperplasia/regeneration from these two studies.
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 and altered renal function (increased urinary
coproporphyrin excretion). 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. The lowest dose of 0.2 mg/kg-day in this study may be a minimal LOAEL for renal
injury.
Application of Uncertainty Factors
A composite uncertainty factor (UF) of 1,000 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; a factor of 3 for the use of a NOAEL that
may be a minimal LOAEL; and a factor of 3 for database deficiencies (lack of a 2-generation
reproductive study).
Calculation of RfD
Using the NOAEL of 0.2 mg/kg-day from the Kociba etal. (1977) and NTP (1991)
studies, the RfD is derived as follows:
RfD =
(0.2 mg/kg-dav) = 2 x 1Q-4 mg/kg-day
1000
where:
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0.2 mg/kg-day = NOAEL, based on the absence of histopathological effects in
kidneys of rats and mice exposed to HCBD in the diet for up to 24
months (Kociba et al., 1977; NTP, 1991).
1,000 = 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; a factor of 3 for
database deficiencies (lack of a two-generation reproductive study);
and a factor of 3 for the use of a NOAEL that may be a minimal
LOAEL.
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 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 8-1.
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 urinary coproporphyrin
excretion and 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 8-2. This pattern is consistent with the hypothesis that renal tumor
formation may require, and be secondary to, renal cytotoxicity induced by exposure !to HCBD.
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Table 8-1. 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 Dose"
(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)1'4
source: Kociba et al. (1977)
Table 8-2. Dose-Related Changes in the Rodent Kidney after Oral Exposure to HCBD,
Chronic Study - Rat (Kociba etal., 1977)
Dose (mg/kg-day)
coproporphyrin
increase
terminal kidney weight
increase (abs. &rel.)
hyperplasia - multi
focal
hyperplasia-
adenomatous
tumors
0.2
• - - . •
-
• -
-
-
2
+
(? only)
- '
?
+
(?only)
. -
20
+
+
+
+
+
8.2.2 Dose-Response Characterization
The Kociba et al. (1977) study was 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
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male rats 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)20. The resulting cancer slope factor is 7.8 x W2 (mg/kg-day)'1. This slope factor
corresponds to a drinking water unit risk of 2.2 x W6 per ng/L (U.S. EPA, 1997), and the
drinking water concentration that corresponds to a lifetime excess cancer risk of 1 x 10'6 is 0.5
jig/L.
1996 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. Under the proposed
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 DMA-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 tumorogenesis. 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 to a wide range of results, and there is currently
no basis to choose among them. In these cases, a margin of exposure analysis is used to evaluate
concern for levels of exposure.
Because both linear (mutagenic) and nonlinear (toxicity associated) mode of action for
carcinogenicity of HCBD may be operating in vivo, both of these approaches have been evaluated
in the draft Ambient Water Quality Criteria for hexachlorobutadiene (U.S. EPA, 1998a) for
characterizing the carcinogenic hazard of HCBD, as discussed below. ;
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 point-of-departure employs the
human equivalent dose. The dose that causes a 10% increase in extra risk is referred to as the
EDj0. The point-of-departure is defined as the 95% lower confidence limit on the ED10, and is •
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referred to as the LED10. The second step in the process islinear extrapolation of the dose-
response .curve from the LED10 to the origin.
The LED10 for HCBD was calculated by fitting the quantal polynomial model2 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 8-1. The EDW 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 * 10"2
(mg/kg-day)"1.
Non-linear Margin of Exposure (MOE) Approach
The MOE approach is used when available data indicate that the dose-response curve for
tumor induction is nonlinear, 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 is
0.30
0.25 -
0.20-
»
0.15-
0.10 -
0.05 -
0.00
Lower confidence
limit on ED10
(LED10)
2 3 4 5
Human Equivalent Dose (mg/kg-day)
Figure 8-1. Renal Tumor Dose Response Curves
2 This modeling was carried out using the Global 86 multistage model software.
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not considered suitable for dose-response extrapolation, and the MOE approach should be
evaluated.
For HCBD, mode-of-action considerations suggest carcinogenicity is secondary to renal
toxicity, which has a threshold (see Section 7.4.3, Mode of Action), and the MOE analysis
becomes an RfD derivation in accordance with the proposed cancer guidelines (U.S. EPA, 1996).
The point-of-departure (Pdp) selected for use was the NOAEL for renal tubular histological
lesions in female rats. 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, 1996). The NOAEL for renal tubular damage was selected rather than an LEDW for
renal damage because quantitative data on incidence and severity of renal non-cancer
histopathological changes were not reported.
The NOAEL for renal damage reported by Kociba et al. was 0.2 mg/kg-day in female rats.
The human equivalent dose for the NOAEL was calculated to be 0.054 mg/kg-day, using the new
scaling factor of body weight raised to the 3/4 power. Based on this, the adjusted point-of-
departure is 0.054 mg/kg-day.
An advisory MOE (which is the composite uncertainty factor for HCBD) is calculated
from appropriate uncertainty factors. The advisory MOE includes: a factor of 10 for! protection
of sensitive human subpopulations; a factor of 3 for extrapolation from animals to humans (since
human equivalent dose is used as the Pdp); a factor of 3 for database deficiency (lack of a 2-
generation reproductive study); and a factor of 3 for use of a NOAEL that may be a minimal
effect LOAEL. The advisory MOE is calculated as follows:
Advisory MOE
10-3-3-3 = 300(rounded)
Applying this advisory MOE to the Pdp (NOAEL) of 0.054 mg/kg-day yields a dose of
2 x 10"4 mg/kg-day, which is equal to the oral RfD. Based on this, the RfD is considered to be
protective not only for non-cancer but also for cancer effects from ingestion of HCBD.
8.23 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 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. :
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Although HCBD metabolites have some indication of mutagenic activity, they are judged
not to play a significant role in tumor causation due to their weak activity. Moreover, mutations
of mitochondrial DNA may result in mitochondria! dysfunction (See Sectioin 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
man 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 MOE approach may be
more appropriate for HCBD. The draft Ambient Water Quality Criteria Document for
Hexachlorobutadiene (U.S. EPA, 1998a) has also recommended using the nonlinear approach for
carcinogenicity assessment of HCBD.
8.2.4 Cancer Potency and Unit Risk
Table 8-3 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 (nag/kg-day)'1. This value is about half of the slope factor of 7.8 x iQr2 (mg/kg-
day)-1 derived previously using the linearized multistage (LMS) model (U.S. EPA, 1997), 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 IQr6 per (ng/L) and the drinking water
concentration that corresponds to a lifetime excess risk of 1 x 10"6 is 0.9 jig/L.
Table 8-3. Summary of Cancer Risk Values for HCBD
Approach
LMSa
(U.S. EPA,
1991)
LEDjo, linear
extrapolation
Nonlinear
Parameter
Slope
Unit Risk
RiskoflxlO-6
LEDjo (tumors)b
Slope
Unit Risk
Risk of 1 xlO'6
Pdp (NOAEL)
Advisory MOE
Value
7.8 x lO'2 (mg/kg-day)-1
2.2 x 10'6 per (|ig/L)
0.5ug/L
2.5 mg/kg-day
4 x 10~2 (mg/kg-day)-1
1.1 x 10-6 per(|ig/L)
0.9[ig/L
0.054 mg/kg-day
300
Animal to human dose extrapolation based on body weight20
b Animal to human dose extrapolation based on body weight3'4
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8.2.5 Discussion of Confidence
The available database associating HCBD and carcinogenicity 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 method which is based on only one data point at a high-dose exceeding the MTD.
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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, 1997). After review of and
response to comments, the final CCL was published on March 2,1998 (63FR 10273, U.S. EPA
1998). 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,
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In the sole judgment of the administrator, regulation of such contaminant presents a
meaningful opportunity for health risk reduction for persons served by public jwater
systems.
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
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•' *f-
present dose-response information, where applicable, for fEreshold and non-threshold health
effects.
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 2 * 10"4
mg/kg-day (U.S. EPA, 1998a). Limited evidence of carcinogenic potential in rodents suggests
that HCBD may be carcinogenic secondary to renal tubular epithelial cell cytotoxiciry. However,
data in humans are lacking. Using the non-linear margin of exposure approach, the RfD is also the
acceptable dose for protection of potential carcinogenic effect.
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 or cytogenetic 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 over time include changes in kidney weight, increased urinary excretion of
coproporphyrin, and increased .renal tubular epithelial hyperplasia.
Othe noncancer effects associated with HCBD exposure in animals include developmental
effects and neurotoxicity (Harleman and Seinen, 1979; Badeva, 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 Semen, 1979).
Results from mutageniciry 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% decrease in body weight and severe renal toxicity. Because
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tihese 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, mitochondria DNA), and the resultant
mitochondria! dysfunction is believed to contribute to the renal cytotoxicity and tumors observed
in animals. Recent evidence suggests that HCBD-induced «2(1-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,
cc^-globulin accumulation cannot be the sole mechanism for HCBD-induced carcinogenesis.
One important issue hi 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 newboms 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).
9.23 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 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 then- 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
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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
Kocibaetal.
(1977)
Jonker et al.
(1993b)
Harleman
and Seinen
(1979)
Stottetal.
(1981)
NIP (1991)
Rat
Sprague-
Dawley
Rat
Wistar
Rat
Wistar
Rat
Sprague-
Dawley
Mouse
B6C3F,
4F
5M
5F
6M
6F
5M
5M
5F
0
1
3
10
30
65
100 .
0
2.25
8
28
0
4.6
14.0
35.3
0
0.2
20
OM OF
3M 5F
12 M 16 F
40M49F
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.
Subchronic Studies
NTP (1991)
Mouse
B6C3F,
10
M
10 F
OM OF
0.1 M 0.2 F
0.4 M 0.5 F
1.5 M 1.8 F
4.9 M 4.5 F
16,8 M 19,2 F
13 weeks
1.5M
0.2 F
4.9 M
0.5 F
Renal tubular cell
regeneration
(increased
epithelial nuclei
and basophilic
staining)
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Table 9-1 (continued)
Study
Species
No./
Sex
Doses
mg/kg-day
Duration
NOAEL
mg/kg-day
LOAEL
mg/kg-day
Effects
Chronic Studies
Kociba et al.
(1977)
Rat
Sprague-
Dawley
39-
40
M,F
0
0.2
2
20
22-24
months
0.2
2
Increased kidney
weight; renal
tubular epithelial
hyperplasia and
neoplasia.
M = male; F = female
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 B6C3F, 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 arid strain used, the length of exposure, amd the
method of administration.
la a subchronic oral exposure study of HCBD in B6C3Fi 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
increase in renal tubular epithelial cell hyperplasia/regeneration and altered renal function
(increased urinary coproporphyrin excretion). The value of this NOAEL from a chronic study is
the same as the equivocal NOAEL of 0.2 rng/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 2 x 10"4 mg/kg-day (U.S. EPA, 1998a). 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 NOAEL of 0.2 mg/kg-day for renal tubular epithelial cell hyperplasia/regeneration from
the Kociba et al.(1977) and NTP (1991) studies. A composite uncertainty factor of 1*000 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); use of a NOAEL that may be closer to
a LOAEL (factor of 3); and database deficiency (factor of 3) because of lack of a 2-generatton
reproductive study.
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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 hi this exposed group. The shape of the dose-response curve
cannot be determined from this data set.
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 1996 proposed Guidelines for Carcinogen Risk Assessment
(USPEA, 1996), HCBD is classified as likely to be carcinogenic to humans. 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 point-of-
departure selected for use was the NOAEL of 0.2 mg/kg-day for renal tubular damage. The
adjusted point of departure of 0.054 mg/kg-day is the human equivalent dose for the NOAEL of
0.2 mg/kg-day. Applying a margin of exposure (MOE) of 300 (which consists of uncertainty
factors of 10 for protection of sensitive human subpopulations, 3 for extrapolation from animals
to human, 3 for data base insufficiency, and 3 for use of a NOAEL that may be a minimal effect
LOAEL) to the adjusted point-of-departure (0.054 mg/kg-day), the resulting dose is 2 x 10"4
mg/kg-day, which is the same as the RfD. 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. Therefore, the non-linear margin of exposure approach may 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.
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
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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). For noncarcinogens, the
estimated HRL level was calculated from the RED 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 mat 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 ranging 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 from Round 2 were collected during the period 1993 to 1997. E.ound
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 jig/L and 10 jig/L, respectively.
The median and 99th percentile for Round 2 detections were 0.30 ng/L and 1.5 ng/L, respectively.
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When monitoring results were comparfd 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 and Round 2 monitoring were 0.114% (74 systems)
and 0.018% (11 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. Date 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.
93.3 Use and Fate Data
Significant quantities of HCBD are generated hi 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).
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 hi water range from 3 to 30 days in rivers and from 30 to 300
days in lakes and ground water (Zoeteman et al., 1980). It is expected that airborne HCBD, like
the structurally similar compound tetrachloroethylene, will react with atmospheric ozone and
hydroxyl radicals, leading to degradation (Atkinson and Carter,"1984; Atkinson, 1987). Estimates
of the atmospheric half-life of HCBD based on this assumption range from 60 days to 1.6 years
(HSDB, 2000).
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HCBD may readily partition from the water into biological tissues, as suggested by its
high log octanolrwater 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, m making a preliminary 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 prehminary 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 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
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population data from the Water Industry Baseline HandbSbk (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 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 ug/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 10'5 mg/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
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estimated akrdrinking 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"
Medium
Drinking Waterb
Food
Air
Adult (ng/kg-day)
8.6
0.15
120
Child (ng/kg-day) :
30
0.24
630
1 See Chapter 5 for derivation of intakes from media other than water
bBased on half the median values for detected hexachlorobutadiene concentrations in Round 1
and Round 2
Table 9-4. Ratios' of Exposures from Various Media to Exposures from Drinking
Water
Exposure Ratio
Food:Drinking Water
AinDrinking Water
Adult
0.02
14
Child
o.oos ;
21
• 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).
Calculation of medium-specific exposure ratios (Table 9-4) indicates that HCBD intake
from air is about 14- 20 fold greater than intake from water. -Therefore, regulation of HCBD hi
drinking water would be unlikely to significantly reduce the risk to sensitive populations.
9.5 Regulatory Determination Summary
i
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
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reasons, EPA may not propose to regulate HCfiD with a ISfPDWR. All final determinations and
future analysis will be presented in the Federal Register Notice covering CCL proposals. •
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External Review Draft—Hexachlorobutadiene—April 2002
10-11
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External Review Draft—Hexachlorobutadiene—April 2002
10-12
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External Review Draft—Hexachlorobutadiene—April 2002
10-13
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External Review Draft—Hexachlorobutadiene—April 2002
10-14
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External Review Draft—Hexachlorobutadiene—April 2002
10-15
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External Review Draft—Hexachlorobutadiene—April 2002
10-16
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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 HI
SDWA
SDWIS
SDWIS FED
SOC
STORET
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 Drag 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
External Review Draft—Hexachlorobutadiene—April 2002
A-l
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APPENDIX B: Round 1 and Round 2 Occurrence
Data Tables for Hexachlorobutadiene
Hexachlorobutadiene Occurrence in Public Water Systems in Round 1 , UGM
1987) results
STATE
TOTAL
UNIQUE PWS
#GWPWS
#SWPWS
%PWS
with detections
%GWPWS
with
detections
%SWPWS
with
detections
% PWS
>HRL
%GWPWS
>HRL
/eSWPWS
>HRL
99% VALUE
j.§jy%
.1-4^
OJ3B
AR
w
448
47
206
3S>
<*s» B 00
CO
DC
)E
I"
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.64
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.50
10
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.50
O.'00%
8.71%
A*
MS
.00%
3BL
S24
A
MA
13
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
0.50
.936
3
0.00%
*?' 0.5C
Ml
SSL
ijo_
HS
1553.
..1.523
0:OQ%
'MM
...' fUDMt
000%
85
71
14
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
< 20.00
2.
'are
,0.00%
...<' ...6.86
NE
NH
'0.25%,
sac
65S
0.861?
NV
NY
356
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
252
123
0.28%
0.40%
0.00%
0.28%
0.40%
0.00%
0.20
5.00
at
0^08%
2.00
36 -
'- .33%
0.00%
- <
UL
•AM*
TX
100.00%
100.00%
0.00%
100.00%
100.00%
0.00%
8.0(
*H
',0.00%
VI
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
••<- 1.00
VT
sr'
sK
4.00
jit
jjg
OOD%
TOTAL
12.768
11,332
1.53E
0.36%
0.32%
0.65%
0.12%
0.07%
0.46%
5.00
External Review Draft—Hexachlorobutadiene—April 2002
B-l
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Appendix B, Round 1 Data (continued)
Hexachlorobutadiene Occurrence in Public Water Systems in Round 1, UCM
(1987) results
STATE
to ,
«ate»
TOTAL
UNIQUE PWS
12,284
#GWPWS
10,960
#SWPWS
1,385
% PWS
with detections
0.35&
%GWPWS
with
detections
'-, '0,30%
%SWPWS
with
detections
' . 0.72%:
* < ;«e//
% PWS
>HRL.
'.004%
' jS*y*/
%GWPWS
>HRL
/ , pp>
'"•** 4 f
%SWPWS
>HRL
\oMt%
* r*'"
99% VALUE
(jiug/L)
T* 6.QC
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 revie
The Health Reference Level used for hexachlorobutadiene is 0.9 uug/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, > 1/2 Health Reference Level, > Health Reference Level = percent of the total number of public water systems with at least one analytic
result that exceeded the MRL, 1/2 Health Reference Level, Health Reference Level, respectively
99th Perccntile Concentration = the concentration value of the 99th percentile of all analytical results (in uug/L)
Median Concentration of Detections = the median analytical value of all the detections (analytical results greater than the MRL) (in uug/L) .
The highlighted states are part of the URCIS (Round 1) 24 State Cross-Section.
External Review Draft—Hexachlorobutadiene — April 2002
B-2
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Appendix B, Round 2 Data
Hexachlorobutadiene Occurrence in Public Water Systems in Round 2, UCM
STATE
TOTAL
UNIQUE PWS
#GWPWS
#SWPWS
% PWS
with
GWPWS
With
SWPWS
with
%PWS
>HRL
GWPWS
>HRL
SWPWS
>HRL
)%VALUI
21
0.00%
0.00
0.00
481
3.36
60
0.00
0.00%
0.00%
11
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.000
107
0.00
0.00
0.00
0.00
0.00
200
.310
1241
6
0.00
0.00%
0.00%
0.00
0.000
. 0.00
as
82
4*79%
^ 0.00%
OJD
0.00
i
tf.00%
8.60
0,00
100.00
0.00
0.00%
0.6
XJ.00%
0.00
0.00
0,00
20
' 014
-ftp
•
M
-0.00
4M
-.is
0.06
tis
237
21
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
1148
191
18
0.00
0.00
0.00
0.00
TAL
24.81
22.29
2.52
0.17
0.13
0.56
0.12
H*vhss».»;s;
lifl
t •','• •s
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 revie
The Health Reference Level used for hexachlorobutadiene is 0.9 nng/L. This is a draft value for working review only.
Total Number of PWSs = the total number of pubHc water systems with records for hexachlorobutadiene
% PWS with detections, > 1/2 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, 1/2 Health Reference Level, Health Reference Level, respectively
99th Peroentile Concentration = the concentration value of the 99th percentile of all analytical results (in ung/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 SDWIS/FED (Round 2) 20 State Cross-Section.
External Review Draft—Hexachlorobutadiene — April 2002
B-3
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